Environmental Biotechnology lecture note-220110

LECTURE NOTE

ENVIRONMENTAL BIOTECHNOLOGY

Dr. Tran Thi My Dieu

Ho Chi Minh, January 2010

ENVIRONMENTAL BIOTECHNOLOGY

Dr. Tran Thi My Dieu

Jan. 2010VAN LANG UNIVERSITY DENTEMA

Chapter 1Chapter 1

INTRODUCTIONINTRODUCTION

VAN LANG UNIVERSITY DENTEMA

CONTENTSCONTENTS

The roles of bioprocesses in The roles of bioprocesses in environmental technologyenvironmental technology

Applications of bioprocesses in Applications of bioprocesses in environmental technologyenvironmental technology


THE ROLES OF BIOPROCESSESTHE ROLES OF BIOPROCESSES

Is it important? Why? Is it important? Why?


APPLICATIONS OF BIOPROCESSES APPLICATIONS OF BIOPROCESSES IN ENVIRONMENTAL TECHNOLOGYIN ENVIRONMENTAL TECHNOLOGY

Biological wastewater treatment Biological wastewater treatment and reuseand reuse

Biological solid waste treatment Biological solid waste treatment and reuseand reuse

Biological air pollution treatment Biological air pollution treatment and reuseand reuse



Biological Biological WWTWWT

UASB

Lan can baûo veä

OÁng thu nöôùc sau xöû lyù

Saøn coâng taùc

Maùng thu nöôùcdaïng raêng cöa

Thieát bò taùch phakhí – loûng - raén

Vaùch höôùngdoøng hình coân

Caàu thang

Voû thieát bò

Hoãn hôïpnöôùc thaûi

Lôùp buøn kî khí

OÁng bôm nöôùcvaøo thieát bò UASB

Boä phaän phaân phoáiñeàu löu löôïng nöôùcthaûi

OÁng thoaùtkhí

Bình haápthuï khí

Boït khí

Dung dòchNaOH 5%

OÁng daãn khí


APPLICATIONS OF BIOPROCESSES APPLICATIONS OF BIOPROCESSES IN ENVIRONMENTAL TECHNOLOGYIN ENVIRONMENTAL TECHNOLOGYBiological wastewater treatment & reuseBiological wastewater treatment & reuse

Granular sludge



India

UASB


A B C D E

A DB EC

752 days

692 days

EA DB C

832 days

A shift in population is noticeable (Dominated by A shift in population is noticeable (Dominated by MethanosaetaMethanosaeta like organism)like organism)-- 832 832 daysdaysGranules Granules –– B reactor ( Bulk B reactor ( Bulk reethareetha seed ) showed increased seed ) showed increased methongenicmethongenicpopulation, Which is reflected in the IAS studies for the same ppopulation, Which is reflected in the IAS studies for the same perioderiod

852

752

832



Biological wastewater treatment & reuseBiological wastewater treatment & reuse

UASB



Suspended growth AS Attached growth AS




400m3/D400m3/D




Trickling filter



RBC



RBC

Nitrogen removal






Biodegradable constituents removalBiodegradable constituents removal

Nutrient removal (N & P)Nutrient removal (N & P)

Specific trace organic compound removalSpecific trace organic compound removal

WHY?WHY?


APPLICATIONS OF BIOPROCESSES APPLICATIONS OF BIOPROCESSES IN ENVIRONMENTAL TECHNOLOGYIN ENVIRONMENTAL TECHNOLOGYBiological solid waste treatment & reuseBiological solid waste treatment & reuse

BTA PROCESS

DigesterDigester



DigesterDigester





Biological solid waste treatmentBiological solid waste treatment

DRANCO PROCESS, BELGIUM, 1984



composting plantcomposting plant

Open heap compostingOpen heap composting

Biological solid waste treatment & reuseBiological solid waste treatment & reuse



Box/container compostingBox/container composting

Drum, with artificial Drum, with artificial aeration systemaeration system




Anaerobic digesterAnaerobic digester

Anaerobic digestion plantAnaerobic digestion plant








Reuse Reuse biowastebiowaste to produce compostto produce compost

Reuse Reuse biowastebiowaste to produce biogasto produce biogas

Increase biodegradable efficiency of Increase biodegradable efficiency of organic materials in landfillorganic materials in landfill

WHY?WHY?




Odor removal by Odor removal by EMEM

MethanMethan gas (landfill gas) removalgas (landfill gas) removal

WHY?WHY?



THE ROLES OF BIOPROCESSESTHE ROLES OF BIOPROCESSES

Environmental friendly treatment Environmental friendly treatment technologies of biodegradable wastes technologies of biodegradable wastes

Reduce harm to environment & public Reduce harm to environment & public health health

Reduce treatment costReduce treatment cost

To transform biodegradable constituents To transform biodegradable constituents into acceptable end productsinto acceptable end products


MarketMarketSolid wastesSolid wastes WastewaterWastewater

BreedingBreeding

Livestock feed Livestock feed processingprocessing

CompostingComposting WWTSWWTS

Fish culturingFish culturing

ResourceResource

Gas recoveryGas recovery

Cassava cultivationCassava cultivation

Cassava harvestingCassava harvesting


QUESTION?QUESTION?

Your Roles?Your Roles?

CHAPTER 2

FUNDAMENTALS OF BIOLOGICAL WASTEWATER TREATMENT

2.1 OVERVIEW OF BIOLOGICAL WASTEWATER TREATMENT 2.1.1 Objectives of biological treatment The overall objectives of the biological treatment of domestic wastewater are to (1) transform dissolved and particulate biodegradable constituents into acceptable end products, (2) capture and incorporate suspended and nonsettleable colloidal solids into a biological flocs or biofilm, (3) transform or remove nutrients such as nitrogen and phosphorus, and (4) in some cases, remove specific trace organic constituents and compounds. For industrial wastewater, the objective is to remove or reduce the concentration of organic and inorganic compounds. Because some of the constituents and compounds found in industrial wastewater are toxic to microorganisms, pretreatment may be required before the industrial wastewater can be discharged to a municipal collection system. For agricultural irrigation return wastewater, the objective is to remove nutrients, specifically nitrogen and phosphorus that are capable of stimulating the growth of aquatic plants. 2.1.2 Role of microorganisms in wastewater treatment The removal of dissolved and particulate carbonaceous BOD and the stabilization of organic matter found in wastewater is accomplished biologically using a variety of microorganisms, principally bacteria. Microorganisms are used to oxidize (or convert) the dissolved and particulate carbonaceous organic matter into simple end products and additional biomass, as represented by the following equation for the aerobic biological oxidation of organic matter: Organic material + xO2 + yNH3 + zPO4

3- microorganisms new cells + tCO2 + uH2O (2-1) where x, y, z, t, u = the stoichiometric coefficient. In the Eq (2-1), oxygen, ammonia and phosphate are used to represent the nutrients needed for the conversion of the organic matter to simple end products (carbon dioxide and water). The term shown over the directional arrow is used to denote the fact that microorganisms are needed to carry out the oxidation process. The term new cells is used to represent the biomass produced as a results of the oxidation of the organic matter. Microorganisms are also used to remove nitrogen and phosphorus in wastewater treatment processes. Specific bacteria are capable of oxidizing ammonia (nitrification) to nitrite and nitrate, while other bacteria can reduce the oxidized nitrogen to gaseous nitrogen. For phosphorus removal, biological processes are configured to encourage the growth of bacteria with the ability to take up and store large amounts of inorganic phosphorus. Because the biomass has a specific gravity slightly greater than that of water, the biomass can be removed from the treated liquid by gravity settling. It is important to note that unless the biomass produced from the organic matter is removed on a periodic basis, complete treatment has not

Environmental Biotechnology Handout ………………………………………………………………………………………………………………………………………………………….

2

been accomplished because the biomass, which itself is organic, will be measured as BOD in the effluent. Without the removal of biomass from the treated liquid, the only treatment achieved is that associated with the bacterial oxidation of a portion of the organic matter originally present. 2.1.3 Types of biological processes for wastewater treatment The principal biological processes used for wastewater treatment can be divided into two main categories: suspended growth and attached growth (or biofilm) processes. Suspended growth processes In suspended growth processes, the microorganisms responsible for treatment are maintained in liquid suspension by appropriate mixing methods. Many suspended growth processes used in municipal and industrial wastewater treatment are operated with a positive dissolved oxygen concentration (aerobic), but applications exist where suspended growth anaerobic (no oxygen present) reactors are used, such as for high organic concentration industrial wastewaters and organic sludges. The most common suspended growth process used for municipal wastewater treatment is the activated sludge process. The activated sludge process was developed around 1913 Lawrence Experiment Station in Massachusetts by Clark and Gage (Metcalf and Eddy, 1930), and by Arden and Lockett (1914) at the Manchester Sewage Works in Manchester, England. The activated sludge process was so named because it involved the production of an activated mass of microorganisms capable of stabilizing a waste under aerobic conditions. In the aeration tank, contact time is provided for mixing and aerating influent wastewater with the microbial suspension, generally referred to as the mixed liquor suspended solids (MLSS) or mixed liquor volatile suspended solids (MLVSS). Mechanical equipment is used to provide the mixing and transfer of oxygen into the process. The mixed liquor then flows to a clarifier where the microbial suspension is settled and thickened. The settled biomass, described as activated sludge because of the presence of active microorganisms, is returned to the aeration tank to continue biodegradation of the influent organic material. A portion of the thickened solids is removed daily or periodically as the process produces excess biomass that would accumulate along with the non-biodegradable solids contained in the influent wastewater. If the accumulated solids are not removed, they will eventually find their way to the system effluent. An important feature of the activated sludge process is the formation of floc particles, ranging in size from 50 to 200 µm, which can be removed by gravity settling, leaving a relatively clear liquid as the treatment effluent. Typically, greater than 99 percent of the suspended solids can be removed in the clarification step. Attached growth processes In attached growth processes, the microorganisms responsible for the conversion of organic material or nutrients are attached to an inert packing material. The organic material and nutrients are removed from the wastewater flowing past the attached growth also known as a biofilm. Packing materials used in attached growth processes include rock, gravel, slag, sand, redwood, and a wide range of plastic and other synthetic materials. Attached growth processes can also be operated as aerobic or anaerobic processes. The packing can be submerged completely in liquid or not submerged, with air or gas space above biofilm liquid layer. The most common aerobic attached growth process used is the trickling filter in which wastewater is distributed over the top area of a vessel containing non-submerged packing


3

material. Historically, rock was used most commonly as the packing material for trickling filters, typical depths ranging from 1.25 to 2.0 m. Most modern trickling filters vary in height from 5 to 10 m and are filled with a plastic packing material for biofim attachment. The plastic packing material is designed such that about 90 to 95 percent of the volume in the tower consists of void space. Air circulation in the void space, by either natural draft or blowers, provides oxygen for the microorganisms growing as an attached biofilm. Influents wastewater is distributed over the packing and flows as a nonuniform liquid film over the attached biofilm. Excess biomass sloughs from the attached growth periodically and clarification is required for liquid/solids separation to provide an effluent with an acceptable suspended solids concentration. The solids are collected at the bottom of the clarifier and removed for waste-sludge processing.

2.2 COMPOSITION AND CLASSIFICATION OF MICROORGANISMS Biological processes for wastewater treatment consist of mixed communities with a wide variety of microorganisms, including bacteria, protozoa, fungi, rotifers and possibly algae. To provide a basic understanding of the nature of microorganisms, the topics introduced in this section are: (1) cell composition and (2) environmental factors affect microbial activity. 2.2.1 Cell composition To support microbial growth in biological system, appropriate nutrients must be available. Reviewing the composition of a typical microbial cell will provide a basis for understanding the nutrients needed for growth. Prokaryotes are composed of about 80 percent water and 20 percent dry material, of which 90 percent is organic and 10 percent is inorganic. The most widely used empirical formula for the organic fraction of cells is C5H7O2N first proposed by Hoover and Porges (1952). About 53 percent by weight of the organic fraction is carbon. The formulation C60H87O23N12P can be used when phosphorus is also considered. It should be noted that both formulation are approximations and may vary with time and species, but they are used for practical purposes. Nitrogen and phosphorus are considered macronutrients because they are required in comparatively large amounts. Prokaryotes also require trace amount of metallic ions, or micronutrients, such as zinc, manganese, copper, molybdenum, iron and cobalt. Because all of these elements and compounds must be derived from the environment, a shortage of any of these substances would limit and in some case, alter growth. 2.2.2 Environmental factors affect microbial activity Environmental conditions of temperature and pH have an important effect on the selection, survival, and growth of microorganisms. In general, optimal growth for a particulate microorganism occurs within a fairly narrow range of temperature and pH, although most microorganisms can survive within much broader limits. Temperatures below the optimum typically have a more significant effect on growth rate than temperature above the optimum; it has been observed that growth rate double with approximately every 10oC increase in temperature until the optimum temperature is reached. According to the temperature range in which they function best, bacteria may be classified a psychophilic, mesophilic or thermophilic. Table 2.1 Temperature classification of biological processes

Type Temperature range (oC) Optimum range (oC) Psychrophilic 10-30 12-18 Mesophilic 20-50 25-40 Thermalphilic 35-75 55-65

Source: Metcalf and Eddy, 2003.


4

The pH of the environment is also a key factor in the growth of organisms. Most bacteria cannot tolerate pH levels above 9.5 or below 4.0. Generally, the optimum pH for bacterial growth lies between 6.5 and 7.5. Different archaea are able to grow at thermophilic and ultrathermophilic (60 to 80oC) temperatures, extremely low pH and high salinity. 2.3 INTRODUCTION TO MICROBIAL METABOLISM Basic to the design of a biological treatment process, or to the selection of the type of biological process to be used, is an understanding of the biochemical activities of microorganisms. Different microorganisms can use a wide range of electron acceptors, including oxygen, nitrite, nitrate, iron, sulfate, organic compounds and carbon dioxide. 2.3.1 Carbon and energy sources for microbial growth To continue to reproduce and function properly, an organism must have sources of energy, carbon for the synthesis of new cellular material, and inorganic elements (nutrients) such as nitrogen, phosphorus, sulfur, potassium, calcium, and magnesium. Organic nutrients (growth factors) may also be required for cell synthesis. Carbon and energy sources, usually referred to as substrates, and nutrient and growth factor requirements for various types of organisms are considered in the following discussions. Carbon sources. Microorganisms obtain their carbon for cell growth from with either organic matter or carbon dioxide. Organisms that use organic carbon for the formation of new biomass are called heterotrophs, while organisms that derive cell carbon from carbon dioxide are called autotrophs. The conversion of carbon dioxide to cellular carbon compounds requires a reductive process, which requires a net input of energy. Autotrophic organisms must therefore spend more of their energy for synthesis than do heterotrophts, resulting in generally lower yields of cell mass and growth rates. Energy sources. The energy needed for cell synthesis may be supplied by light or by a chemical oxidation reaction. Bacteria can oxidize organic or inorganic compounds to gain energy. Those organisms that are able to use light as an energy sources are called phototrophs. Phototrophic organism may be either heterotrophic (certain sulfur-reducing bacteria) or autotrophic (algae and photosynthetic bacteria). Organisms phototrophs, chemotrophs may be either heterotrophic (protozoa, fungi, and most bacteria) or autotrophic (nitrifying bacteria). Chemoautotrophs obtain energy from the oxidation of reduced inorganic compounds, such as ammonia, nitriate, ferrous iron, and sulfile. Chemoheterotrophs usually derive their energy from the oxidation of organic compounds. The energy-producing chemical reactions by chemotrophs are oxidation-reduction reactions that involve the transfer of electrons from an electron donor to an electron acceptor. The electron donor is oxidized and the electron acceptor is reduced. The electron donors and acceptors can be either organic or inorganic compounds, depending on the microorganism. The electron acceptor may be available within the cell during metabolism (endogenous) or it may be obtained from outside the cell (i.e. dissolved oxygen). Organisms that generate energy by enzyme-mediated electron transport to an external electron acceptor are said to have a respiratory metabolism. The use of an internal electron acceptor is termed fermentative metabolism and is less efficient energy-yielding process that respiration. Heterotrophic organisms that are strictly fermentative are characterized by lower growth rates and cell yields that respiratory heterotrophs. When oxygen is used for the electron acceptor the reaction is termed aerobic, and reactions involving other electron acceptors are considered anaerobic. The term anoxic is used to


5

distinguish the use of nitrate or nitrate for electron acceptors from the others under anaerobic conditions. Under anoxic condition, nitrate or nitrate reduction to gaseous nitrogen occurs, and this reaction is also referred to as biological denitrification. Organism that can only meet their energy needs with oxygen are called obligate aerobic microorganisms. Some bacteria can use oxygen or nitrate/nitrite as electron acceptors when oxygen is not available. These bacteria are called facultative aerobic bacteria. Organism that generate energy by fermentation and that can exist only in an environment that is devoid of oxygen are obligate anaerobes. Facultative anaerobes have the ability to grow in either the presence or absence of molecular oxygen and fall into two subgroups, based on their metabolic abilities. True facultative anaerobes can shift from fermentative to aerobic respiratory metabolism, depending on the presence or absence of molecular oxygen. Aerotolerant anaerobes have a strictly fermentative metabolism but are relatively insensitive to the presence of molecular oxygen. 2.3.2 Nutrient and growth factor requirement Nutrient, rather than carbon or energy sources, may at times be the limiting material for microbial cell synthesis and growth. The principal inorganic nutrients needed by microorganisms are N, S, O, K, Mg, Ca, Fe, Na and Cl. Minor nutrients of importance include Zn, Mn, Mo, Se, Co, Cu and Ni. Required organic nutrients, known as growth factor, are compounds needed by an organism as precursors or constituents of organic cell material, which cannot be synthesized from other carbon sources. Although growth factor requirements differ from one organism to another, the major growth factors fall into the following three classes: (1) amino acids, (2) nitrogen bases (i.e. purines and pyrimidines) and (3) vitamins. For municipal wastewater treatment sufficient nutrients are generally present, but for industrial wastewaters nutrients may need to be added to the biological treatment processes. The lack of sufficient nitrogen and phosphorus is common especially in the treatment of food – processing wastewaters or wastewater high in organic content. Using the formula C60H87O23N12P, it is possible to estimate amount of nitrogen and phosphorus needed per 100 g of cell biomass. 2.4 BACTERIAL GROWTH 2.4.1 Bacterial reproduction Bacteria can reproduce by binary fission, by asexual mode or by budding. Generally, they reproduce by binary fission, in which the original cell becomes two new organisms. The time required for each division, which is termed the generation time can vary from days to less than 20 min. This rapid change in biomass with time is a hypothetical example, for in biological treatment systems, bacteria would not continue to divide indefinitely because of environmental limitations, such as substrate and nutrient availability. 2.4.2 Bacterial growth patterns in a batch reactor Bacterial growth in a batch reactor is characterized by identifiable phases as illustrated on Fig. 2.1. The curves shown on Fig. 2.1 represent what occurs in a batch reactor in which, at time zero, substrate and nutrients are present in excess an only a very small populations of biomass exists. As substrate is consume, four distinct growth phases develop sequentially. 1. The lag phase. Upon addition of the biomass, the lag phase represents the time required for

the organism to acclimate to their new environment before significant cell division and


6

biomass production occur. During the lag phase enzyme induction may be occurring and/or the cells may be acclimating to changes in salinity, pH, or temperature. The apparent extent of the lag phase may also be affected by ability to measure the low biomass concentration during the initial batch phase.

2. The exponential-growth phase. During the exponential-growth phase, bacterial cells are multiplying at their maximum rate, as there is no limitation due to substrate or nutrients. The biomass growth curve increase exponentially during this period. With unlimited substrate and nutrients the only factor that affects the rate of exponential growth is temperature.

3. The stationary phase. During this phase, the biomass concentration remains relatively constant with time. In this phase, bacterial growth is no longer exponential and the amount of growth is offset by the death of cells.

4. The death phase. In the death phase, the substrate bas been depleted so that no growth is occurring, and the change in biomass concentration is often observed as an approximate constant fraction of the biomass remaining that is lost each day.

2.4.3 Bacterial growth and biomass yield In biological treatment processes, cell growth occurs concurrent with the oxidation of organic or inorganic compounds, as described above. The ratio of the amount of biomass produced to the amount of substrate consumed (g biomass/g substrate) is defined as the biomass yield, and typically is defined relative to the electron donor used. g biomass produced Biomass yield Y = --------------------------------------------- (2-2) g substrate utilized (i.e. consumed) For example, for aerobic heterotrophic reactions with organic substrates, the yield is expressed as g biomass/g organic substrate; for nitrification the yield is expressed as g biomass/g NH4-N oxidized and for the anaerobic degradation of volatile fatty acids (VFAs) to produce methane, the yield is expressed as g biomass/g VFAs used. Where specific compounds are measured and known such as ammonia, the yield is quantified relative to the amount of compound used. For aerobic or anaerobic treatment of municipal and industrial wastewater containing a large number of organic compounds, the yield is based on a measureable parameter reflecting the overall organic compounds consumption, such as COD or BOD. Thus, the yield would be g biomass/g COD removed or g biomass/g BOD removed. 2.4.4 Measuring biomass growth Because biomass is mostly organic material, an increase in biomass can be measured by volatile suspended solids (VSS) or particulate COD (total COD minus soluble COD). Other parameters that are used to indicate biomass growth are protein content, DNA and ATP, a cellular compound involved in energy transfer. Of these growth measurement parameters, VSS is the parameter used most commonly to follow biomass growth in full-scale biological wastewater treatment systems because its measurement is simple and minimal time is required for analysis. It should be noted that the VSS measured includes other particulate organic matter in addition to biomass. Most wastewater contain some amount of non-biodegradable VSS and possibly influent VSS that may be slowly degraded in the biological reactor. These solids are included with biomass in the VSS measurement. Nevertheless the VSS measurement is used as an apparent indicator of biomass production and also provides as useful measurement of reactor solids in general.


7

For laboratory research on biological treatment processes, growth parameters that can be related to true microbial mass are often used. Of these, protein is the most popular growth parameter due to the relative ease of measurement and the fact that about 50 percent of biomass dry weight is protein. Both ATP and DNA have also been used, especially where the reactor solids contain proteins and other solids that are not associated with biomass. Where very low biomass concentrations are involved, turbidity measurement may be used to provide a rapid and simple means of observing cell growth. Bacterial cell counts have also been used to enumerate the biomass population. A portion of a diluted sample is applied to an agar growth plate, and after incubation, the number of colonies formed are counted and used to determine the number of bacterial cells in the culture. It should be noted, however, that not all bacteria are culturable. 2.4.5 Estimating biomass yield and oxygen requirements As given by Eq. (2-1), a definite stoichiometric relationship exists between the substrate removed, the amount of oxygen consumed during aerobic heterotrophic biodegradation, and the observed biomass yield. The most common approach used to define the fate of the substrate is to prepare a COD mass balance. The COD is used because the substrate concentration in the wastewater can be defined in terms of its oxygen equivalence, which can be accounted for being conserved in the biomass or oxidized. In general, the exact stoichiometry involved in the biological oxidation of a mixture of wastewater compounds is never known. However, for the purpose of illustration, assume organic matter can be represented as C6H12O6 (glucose) and new cells can be represented as C5H7O2N. Thus neglecting nutrients other than nitrogen, Eq. (2-1) can be written as: 3C6H12O6 + 8O2 + 2NH3 2C5H7O2N + 8CO2 + 12H2O (2-3) 3 (180) 8 (32) 2 (17) 2 (113) As given by the above equation, the substrate used (glucose in this case) is divided between that found in new cells and that oxidized. The yield based on the glucose consumed can be obtained as follows:

( )( ) =

⎟⎟⎠

⎞⎜⎜⎝

⎛

⎟⎟⎠

⎞⎜⎜⎝

⎛

=∆∆=

mole

g

mole

g

OHC

NOHCY

1803

1132

6126

275 0.42 g cells/g glucose used

In practice, COD and VSS are used to represent the organic matter and the new cells, respectively. To express the yield on a COD basis, the COD of glucose must be determined. The COD of glucose can be determined by writing a balanced stoichiometric reaction for the oxidation of glucose to carbon dioxide as follows: C6H12O6 + 6O2 6CO2 + 6H2O (2-4) (180) 6 (32)

The COD of glucose is

( )( ) =

⎟⎟⎠

⎞⎜⎜⎝

⎛

⎟⎟⎠

⎞⎜⎜⎝

⎛

=∆

∆=

mole

g

mole

g

OHC

OCOD

180

326

6126

2 1.07 g O2/ g glucose

The theoretical yield expressed in terms of COD, accounting for the portion of the substrate converted to new cells, is


8

( )( ) =

⎟⎠⎞⎜

⎝⎛⎟⎟⎠

⎞⎜⎜⎝

⎛

⎟⎟⎠

⎞⎜⎜⎝

⎛

=∆

∆=

egglugCOD

mole

g

mole

g

asCODOHC

NOHCY

cos07.11803

1132

6126

275 0.39 g cells/g COD used.

It should be noted that the actual observed yield in a biological treatment process will be less than the value given above, because a portion of the substrate incorporated into the cell mass will be oxidized with time by the bacteria to obtain energy for cell maintenance. The quantity of oxygen utilized can be accounted for by considering: (1) the oxygen used for substrate oxidation to CO2 and H2O, (2) the COD of the biomass, and (3) the COD of any substrate not degraded. Based on the formula C5H7O2N, the oxygen equivalent of the biomass (typically measured as VSS) is approximately 1.42 g COD/g biomass VSS, as given below. C5H7O2N + 5O2 5CO2 + NH3 + 2H2O (2-5) (113) 5 (32) The COD of cell tissue is

( )( ) =

⎟⎟⎠

⎞⎜⎜⎝

⎛

⎟⎟⎠

⎞⎜⎜⎝

⎛

=∆

∆=

mole

g

mole

g

NOHC

OCOD

113

325

275

2 1.42 g O2/g cells

Based on the above relationships, the oxygen consumed per unit of COD utilized for the reaction given by Eq. (2-3) can be determined from a mass balance on COD as follows: COD utilized = COD cells + COD of oxidized substrate (2-6) The COD of the oxidized substrate is equal to oxygen consumed; thus: Oxygen consumed = COD utilized – COD cells (2-7) Oxygen consumed = 1.07 g O2/g glucose x 3 mole x 180 g glucose/mole – 1.42 g O2/g cells x 2 moles x 113 g cells/mole = 577.8 – 320.9 = 256.9 g O2 Thus, the oxygen consumed per unit of COD used is: Oxygen consumed 256.9 g O2 ---------------------- =-------------------------------------------------------------- = 0.44 g O2/g COD used Glucose as COD 3 mole (1.07 g COD/g glucose) (180 g glucose/mol) The amount of oxygen required based on the COD balance as given above is in agreement with the oxygen used based on the stoichiometry as defined by Eq. (2-3) in which 8 moles of oxygen are required for 3 mole of glucose. Oxygen used 8 (32 g O2/mole) --------------------- = ----------------------------------------------- = 0.44 g O2/g COD used Glucose as COD 3 (180 g/mol) (1.07 g COD/g glucose)


9

2.4.6 Observed versus synthesis yield In the evaluation and modeling of biological treatment systems as distinction is made between the observed yield and the synthesis yield (or true yield). The observed biomass yield is based on the actual measurements of biomass production and substrate consumption and is actually less than the synthesis yield, because of cell loss concurrent with cell growth. In full-scale wastewater treatment processes the term solid production (or solid yield) is also used to describe the amount of VSS generated in the treatment process. The terms is different from the synthesis biomass yield values because it contains other organic solids from the wastewater that are measured as VSS, but are not biological. The synthesis yield is the amount of biomass produced immediately upon consumption of the growth substrate or oxidation of the electron donor in the case of autotrophic bacteria. The synthesis yield is seldom measure directly and is often interpreted from evaluating biomass production data for reactors operating under different conditions. Synthesis yield values for bacterial growth are affected by the energy that can be derived from the oxidation-reduction reaction, by the growth characteristics of the carbon source, by the nitrogen source, and by environmental factors such as temperature, pH and osmotic pressure. As illustrated in this section, the synthesis yield can be estimated if the stoichiometry or the amount of energy produced in the oxidation-reduction reaction is known. 2.5 MICROBIAL GROWTH KINETICS The performance of biological processes used for wastewater treatment depends on the dynamics of substrate utilization and microbial growth. Effective design and operation of such systems requires an understanding of the biological reactions occurring and an understanding of the basic principles governing the growth of microorganisms. Further, the need to understand all of the environmental conditions that affect the substrate utilization and microbial growth rate cannot be overemphasized, and it may be necessary to control such conditions as pH and nutrients to provide effective treatment. The purpose of this section include (1) microbial growth kinetics terminology, (2) rate of utilization of soluble substrate, (3) other rate expressions for the utilization of soluble substrate, (4) rate of soluble substrate production from biodegradable particulate organic matter, (5) the rate of biomass growth with soluble substrates, (6) kinetic coefficients for substrate utilization and biomass growth, (7) the rate of oxygen uptake, (8) effects of temperature, (9) total volatile suspended solids and active biomass, and (19) net biomass and observed yield. 2.5.1 Microbial growth kinetics terminology The kinetics of microbial growth govern the oxidation (i.e. utilization) of substrate and the production of biomass, which contributes to the total suspended solids concentration in a biological reactor. Because municipal and industrial wastewater contain numerous subsrates, the concentration of organic compounds is defined, most commonly, by the biodegradable COD (bCOD) or UBOD, both of which are comprised of soluble (dissolved), colloidal and particulate biodegradable components. Both bCOD and UBOD represent measurable quantities that apply to all of the compounds. In the formulation of kinetic expressions in this section biodegradable soluble COD (bsCOD) will be used to quantify the fate of biodegradable organic compounds because it easily relates to the stoichiometry of substrate oxidized or used in cell growth. The biomass solids in a bioreactor are commonly measured as total suspended solid (TSS) and volatile suspended solid (VSS). The mixture of solids resulting from combining recycled sludge with influent wastewater in the bioreactor is termed mixed liquor suspended solids (MLSS) and


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mixed liquor volatile suspended solids (MLVSS). The solids are comprised of biomass, non-biodegradable volatile suspended solids (nbVSS), and inert inorganic total suspended solids (iTSS). The nbVSS is derived from the influent wastewater and is also produced as cell debris from endogenous respiration. The ITSS originates in the influent wastewater. 2.5.2 Rate of utilization of soluble substrate The goal in biological wastewater treatment is, in most cases, to deplete the electron donor (i.e. organic compounds in aerobic oxidation). For heterotrophic bacteria the electron donors are the organic substances being degraded; for autotrophic nitrifying bacteria it is ammonia or nitrate or nitrite or other reduced inorganic compounds. The substrate utilization rate in biological systems can be modeled with the following expression for soluble substrates. Because the mass of substrate is decreasing with time due to substrate utilization and Eq. (2-8) is used in substrate mass balances, a negative value is shown.

SK

kXSr

ssu +

−= (2-8)

Where rsu = rate of substrate concentration change due to utilization, g/m3.d k = maximum specific substrate utilization rate, g substrate/g microorganisms.d X = biomass (microorganism) concentration, g/m3 S = growth-limiting substrate concentration in solution, g/m3 Ks = half-velocity constant, substrate concentration at one-half the maximum specific

substrate utilization rate, g/m3 Equation (2-8) will be recognized as a saturation-type equation. For substrate removal, Eq. (2-8) has been referred to as the Michaelis-Menten equation (Bailey and Ollis, 1986), Equation (2-8) is also of the form proposed by Monod for the specific growth rate of bacteria in which the limiting substrate is available to the microorganisms in a dissolved form (Monod, 1942; 1949). When the substrate is being used at its maximum rate, the bacteria are also growing at their maximum rate. The maximum specific growth rate of the bacteria is thus related to the maximum specific substrate utilization rate as follows.

kYm =µ (2-9) and

Yk mµ= (2-10)

Where µm = maximum specific bacterial growth rate, g new cells/g cells.d k = maximum specific substrate utilization rate, g/g.d Y = true yield coefficient, g/g (defined earlier) Using the definition for the maximum specific substrate utilization rate given by Eq. (2-8), the substrate utilization rate is also reported in the literature as:

( )SKY

XSr

s

msu +

−= µ (2-11)


11

2.5.3 Other rate expressions for the utilization of soluble substrate In reviewing kinetic expressions used to describe substrate utilization and biomass growth rates, it is very important to remember that the expression used to model biological processes are all empirical, based on experimentally determined coefficient values. Besides the substrate limited relationship presented above, other expressions that have been used to describe substrate utilization rates include the following: rsu = -k (2-12) rsu = -kS (2-13) rsu = - kXS (2-14) rsu = -kX.(S/So) (2-15) The particular rate expression used to define kinetics of substrate utilization depends mainly on the experimental data available to if the kinetic equations and the application of the kinetic model. In some cases, the first-order model shown in Eq. (2-14) is satisfactory for describing substrate utilization rates when the biological treatment process will be operated at relatively low substrate concentrations. Fundamental in the use of any rate expression is its application in a mass-balance analysis to be discussed in the following section. Also with regard to modeling biological treatment processes, kinetic models should not be applied outside of the range of the conditions used to develop model coefficients. 2.5.4 Rate of soluble substrate production from biodegradable particulate organic matter The rate expressions for substrate utilization and biomass growth presented thus far are based on the utilization of soluble substrates. In municipal wastewater treatment only about 20 to 50 percent of the degradable organic material enters as soluble compounds, and for some industrial wastewaters the soluble organic material may be a low to moderate fraction of the total degradable organic substrates. Bacteria cannot consume the particulate substrates directly and employ extracellular enzymes to hydrolyze the particulate organics to soluble substrates, which can be according to the rate expressions described above. The particulate substrate conversion rate is also a rate-limiting process that is dependent on the particulate substrate and biomass concentrations. A rate expression for particulate substrate conversion is shown as follows (Grady et al., 1999):

( )( )X

PK

XXPk

rx

p

psc +−=, (2-16)

Where rsc,P = rate of change of particulate substrate concentration due to conversion to soluble

substrate, g/m3/d KP = maximum specific particulate conversion rate, g P/g X.d

P = particulate substrate concentration, g/m3 X = biomass concentration, g/m3 Kx = half-velocity degradation coefficient, g/g

The particulate degradation concentration is expressed relative to the biomass concentration, because the particulate hydrolysis is related to the relative contact area between the non-soluble organic material and the biomass.


12

2.5.5 Rate of biomass growth with soluble substrates The following relationship between the rate of growth and the rate of substrate utilization is applicable in both batch and continuous culture systems.

XkSK

kXSYXkYrr d

sdsug −

+=−−= (2-17)

Where rg = net biomass production rate, g VSS/m3.d Y = synthesis yield coefficient, g VSS/g bs COD kd = endogenous decay coefficient, g VSS/g VSS.d Other terms are as defined above. If both side of Eq. (2-17) are divided by the biomass concentration X, the specific growth rate is defined as follows:

dS

g kSK

kSY

X

r−

+==µ (2-18)

Where µ = specific biomass growth rate, g VSS/g VSS.d As shown, the specific growth rate corresponds to the change in biomass per day relative to the amount of biomass present, and is a function of the substrate concentration and the endogenous decay coefficient. The endogenous decay coefficient accounts for the loss in cell mass due to oxidation of internal storage products for energy for cell maintenance, cell death, and predation by organisms higher in the food chain. It should be noted that micoorganisms in all growth phases require energy for cell maintenance; however, it appears that the decay coefficient most probably changes with cell age. Usually, these factor are lumped together, and it is assumed that the decrease in cell mass caused by them is proportional to the concentration of organisms present. The decrease in mass is often indentified in the literature as the endogenous decay. In Eq. (2-17) the coefficient kd is the endogenous decay rate coefficient. In biological treatment processes, both the substrate utilization and biomass growth rates are controlled by some limiting substrate, as given by Eqs. (2-11) and (2-17). The growth limiting substrate can be any of the essential requirements for cell growth (i.e., electron donor, electron acceptor, or nutrients), but often it is the electron donor that is limiting, as other requirements are usually available in excess. Thus, when the term substrate is used to described growth kinetics, it generally refers to the electron donor. 2.5.6 Kinetic coefficients for substrate utilization and biomass growth The coefficient values (k, Ks, Y and kd) used to predict the rate of substrate utilization and biomass growth can vary as a function of the wastewater source, microbial population, and temperature. Kinetic coefficient values are determined from bench-scale testing or full-scale plant test results. For municipal and industrial wastewater the coefficient values represent the net effect of microbial kinetics on the simultaneous degradation of variety of different wastewater constituents. Typical kinetics coefficient values are reported in Table 2.2 for the aerobic oxidation of BOD in domestic wastewater.


13

Table 2.2 Typical kinetic coefficients for the activated sludge process for the removal of organic matter from domestic wastewater

Value* Coefficient Unit

Range Typical k g bsCOD/g VSS.d 2-10 5Ks mg/L BOD 25-100 60 mg/L bsCOD 10-60 40Y mg VSS/mg BOD 0.4-0.8 0.6 mg VSS/mg bsCOD 0.3-0.6 0.4kd g VSS/g VSS.d 0.06-0.15 0.10

* values reported are for 20oC Source: Metcalf and Eddy, 2003. 2.5.7 Rate of oxygen uptake The rate of oxygen uptake is related stoichiometrically to the organic utilization rate and growth rate. Thus, the oxygen uptake rate can be defined as: ro = - rsu – 1.42rg (2-19) where ro = oxygen uptake rate, g O2/m3.d rsu = rate of substrate utilization, g bsCOD, m3.d 1.42 = the COD of cell tissue, g bsCOD/g VSS rg = rate of biomass growth, g VSS/m3.d A negative sign is required in front of the term rsu in Eq. (2-19), because the rate of substrate utilization as given by Eq. (2-11) is negative (i.e., the substrate concentration decreases with time). The factor 1.42 represents the COD of cell tissue as defined previously by Eq. (2-5). 2.5.8 Effects of temperature The temperature dependence of the biological reaction-rate constants is very important in assessing the overall efficiency of a biological treatment process. Temperature not only influences the metabolic activities of the microbial population but also has a profound effect on such factors as gas-transfer rates and the settling characteristics of the biological solids. The effect of temperature on the reaction rate of a biological process is expressed as follows:

( )2020

−= TT kk θ (2-20)

Where kT = reaction-rate coefficient at temperature T, oC k20 = reaction-rate coefficient at 20oC; θ = temperature-activity coefficient T = temperature, oC Value for θ in biological systems can vary from 1.02 to 1.25. 2.5.9 Total volatile suspended solids and active biomass The kinetic expressions used to describe biological kinetics and growth rate are related to the active biomass concentration X in the treatment reactor. In reality the VSS in a reactor consists of more than active biomass, and the fraction of active biomass can vary depending on the wastewater characteristics an operating conditions. The other components that contribute to the


14

VSS concentration are cell debris, following endogenous decay, and non-biodegradable VSS (nbVSS) in the influent wastewater fed to the biological reactor. During cell death, cell lysis occurs with the release of cellular materials into the liquid for consumption by other bacteria. A portion of the cell mass (cell wall) is not dissolved and remains as non-biodegradable particulate matter in the system. The remaining non-biodegradable material is referred to as cell debris and represents about 10 to 15 percent of the original cell weight. Cell debris is also measured as VSS and contributes to the total VSS concentration measured in the reactor mixed liquor. The rate of production of cell debris is directly proportional to the endogenous decay rate. rxd = fd.kd.X (2-21) where rxd = rate of cell debris production, g VSS/m3.d fd = fraction of biomass that remains as cell debris, 0.10-0.15 g VSS/g VSS The nbVSS concentration resulting from cell debris is typically a relatively small fraction of the VSS in a bioreactor used to treat municipal and some industrial wastewaters. As noted above, a variable amount of MLVSS that is not biomass originates from the nbVSS in the influent wastewater. For typical untreated municipal wastewater the nbVSS concentration may be in the range from 60 to 100 mg/L, and following primary treatment may range from 10 to 40 mg/L. Total volatile suspended solids. The VSS production rate in the aeration tank can be defined as the sum of the biomass production given by Eq. (2-17), the nbVSS production viven by Eq (2-21) and the nbVSS in he influent: rXT,VSS = -Yrsu – kdX + fdkdX + Q.Xo,i/V (2-22)

where rXT,VSS = total VSS production rate, g/m3.d Q = influent flow-rate, m3/d

Xo.i = influent nbVSS concentration, g/m3 V = volume of reactor, m3 Other terms are as defined respectively.

Active biomass

From Eq. (2-22), the fraction of active biomass in the mixed-liquor VSS (MLVSS) is the ratio of the sum of the growth and decay term divided by the total MLVSS production:

VSSX

dsuactX

Tr

XkYrF

,,

−−= (2-23)

Where Fx, act = active fraction of biomass in MLVSS, g/g. 2.5.10 Net biomass yield and observed yield Net biomass yield and observed yield The term true yield was defined as the amount of biomass produced during cell synthesis relative to the amount of substrate degraded. In the design and analysis of biological treatment processes,

Net nbVSS from soluble

bCOD

nbVSS from cells

nbVSS in influent


15

two other yield terms are important: (1) the net biomass yield and (2) the observed solid yield. The first is used as an estimate of the amount of active microorganisms in the system, and the second as the amount of sludge production. Net biomass yield. The net biomass yield is the ratio of the net biomass growth rate to the substrate utilization rate: Ybio = - rg/rsu (2-24) Where Ybio = net biomass yield, g biomass/g substrate used. Observed yield. The observed yield accounts for the actual solids production that would be measured for the system and is shown as follows: Yobs = - rxT,VSS/rsu (2-25) Where Yobs = observed yield, g VSS produced/g substrate removed 2.6 MODELING SUSPENDED GROWTH TREATMENT PROCESSES 2.6.1 Description of suspended growth treatment processes The complete-mix reactor with recycle will be considered in the following discussion as a model for suspended growth processes. The schematic flow diagram shown on Fig. 2.1 Include the nomenclature used in the following mass-balance equations. A similar complete-mix reactor may be used in laboratory studies to asses wastewater treatability and to obtain model kinetic coeffients. All biological treatment reactor designs are based on using mass balances across a define volume for each specific constituent of interest (i. e. biomass, substrate, etc.). The mass balance includes the flow-rates for the mass of the constituent entering and/or leaving the system and appropriate reaction rate terms for the depletion or production of the constituent within the system. The units for a mass balance are usually given in mass per volume per time. For all mass balance a check of the units is recommended to assure that the mass-balance equations are correct.

Fig. 2.1 Schematic diagram of activated-sludge process with model nomenclature: (1) with wasting from

the sludge return line and (b) with wasting from the aeration tank (Metcalf and Eddy, 2003).


16

2.6.2 Biomass mass balance A mass balance for the mass of microorganisms in the complete-mix reactor shown on Fig. 2.1a can be written as follows: - General word statement: (2-26) - Simplified word statement:

Accumulation = inflow – outflow + net growth (2-27) - Symbolic representation:

( ) VrXQXQQQXVdt

dXgRWeWo ++−−= ][ (2-28)

Where dX/dt = rate of change of biomass concentration in reactor measured as g VSS/m3.d

V = reactor volume (i.e., aeration tank), m3 Q = influent flow-rate, m3/d Xo = concentration of biomass in influent, g VSS/m3 Qw = waste sludge flow-rate, m3/d Xe = concentration of biomass in influent, g VSS/m3 XR = concentration of biomass in return line from clarifier, g VSS/m3 rg = net rate of biomass production, g VSS/m3.d

It it is assumed that the concentration of microorganisms in the influent can be neglected and that steady-state conditions prevail (dX/dt = 0), Eq. (2-29) can be simplified to yield:

(Q – QW)Xe + QwXR = rgV (2-29) If Eq. (2-29) is combined with Eq. (2- 17), the result is: ( )

dsuRWeW kX

rY

VX

XQXQQ −−=+− (2-30)

Where X = concentration of the biomass in the reactor, g/m3. The inverse of the term on the left-hand side of Eq. (2-30) is defined as the average solids retention time (SRT) as given below:

( ) RWeW XQXQQ

VXSRT

+−= (2-31)

Note that the numerator in Eq. (2-31) represents the total mass of solids in the aeration tank and the denominator corresponds to the amount of solids lost per day in the effluent and by

Rate of accumulation of microorganism within the system

boundary

Rate of flow of microorganism into the system

boundary

= -

Rate of flow of microorganism

out of the system boundary

+

Net growth of microorganism

within the boundary


17

intentional wasting (QW). By definition the SRT is the solids in the system divided by the mass of solids removed per day. Using the above definition of STR, Eq. (2-30) can be written as:

dsu kX

rY

SRT−−=1 (2-32)

The term of 1/SRT is also related to µ, the specific biomass growth rate, as given by the Eq. (2-33):

µ=SRT

1 (2-33)

Thus, for a complete-mix activated-sludge process, the SRT (which can be controlled by solids wasting) is the inverse of the average specific rate that relates to the process biokinetics. In Eq. (2-32), the term (-rsu/X) is known as the specific subsrtate utilization rate U. The specfic substrate utilization rate U is calculated as follows:

( )X

SS

VX

SSQ

X

rU oosu

τ−=−== (2-34)

Where U = specific substrate utilization rate, g BOD or g COD/g VSS.d Q = wastewater flow-rate, m3/d So = influent soluble substrate concentration, g BOD or bsCOD/m3 S = effluent soluble substrate concentration, g BOD or bsCOD/m3 V = volume of aeration tank, m3

X = biomass concentration, g/m3 τ = hydraulic detention time, V/Q, d Other terms as defined above

Substituting Eq. (2-8) into Eq. (2-32) yields:

ds

kSK

YkS

SRT−

+=1 (2-35)

The solids retention time (SRT) is an important design and operating parameter for the activated-sludge process. The SRT is the average time the activated-sludge solids are in the system. Assuming that the solids inventory in the clarifier shown on Fig. 2.1a is nefligicle compared to that in the aeration tank, the SRT is determined by dividing the the mass of solids in the aeration tank by the solids removed daily via the effluent and by wasting for process control. For many activated-sludge processes, where good flocculation occurs and the clarifier is designed properly, the effluent VSS is typically less than 15 g/m3. Where the effluent VSS is low, excess solids must be removed from the system by wasting. Wasting is accomplished most commonly by removing biomass (sludge) from the clarifier underflow recycle line as shown as Fig. 2.1a. Alternatively, wasting can be accomplished from the aeration tank as shown on Fig. 2.1b. Solving Eq. (2-35) for the effluent dissolved substrate concentration S yields:

[ ]( ) 1

1−−

×+=d

ds

kYkSRT

SRTkKS (2-36)


18

It should be noted that Eq. (2-36), the effluent soluble substrate concentration for a complete-mix activated-sludge process is only a function of the SRT and kinetic coefficients for growth and decay. The effluent substrate concentration is not related to the influent soluble substrate concentration, but as will be shown in other mass balances, the influent concentration affects the biomass concentration. 2.6.3 Substrate mass balance The mass balance for substrate utilization in the aeration tank (Fig. 2.1a) is: Accumulation = inflow – outflow + generation

VrQSQSVdt

dSsuo +−= (2-37)

Where So = influent soluble substrate concentration, g/m3 Substituting the value for rsu (Eq. 2-8) and assuming steady-state condition (dS/dt = 0), Eq. (2-37) can be rewritten as:

⎟⎟⎠

⎞⎜⎜⎝

⎛+⎟⎟

⎠

⎞⎜⎜⎝

⎛=−

SK

kXS

Q

VSS

so (2-38)

The volume of the aeration tank divided by the influent flow-rate is τ, the hydraulic retention time. If Eq. (2-35) is solved for the term S/(Ks + S) and substituted into Eq. (2-38), the following expression is obtained for the biomass concentration in the aeration tank:

( )⎥⎦

⎤⎢⎣

⎡+

−⎟⎠⎞

⎜⎝⎛=

SRTk

SSYSRTX

d

o

1τ (2-39)

As given by Eq. (2-39), the reactor biomass concentration is a function of the system SRT, the aerobic tank τ, the synthesis yield coefficient, the amount of substrate removed (So –S), and the endogenous decay coefficient. The same equations can be applied to describe an activate-sludge process with no clarifier and thus no return sludge flow. For the case with no return sludge, all of the solid produced are present in the effluent from the aeration tank and the SRT equals the τ.

τ==QX

VXSRT (2-40)

The importance of the system SRT in determining the effluent soluble substrate concentration and aeration tank biomass concentration is clear from an examination of Eqs. (2-36) and (2-39). As indicated by the waste sludge flow term (Qw) is Eq. (2-31), a selected SRT value can be maintained by the amount of solids wasted per day to control the process performance. Similarly, it can be shown that by wasting from the aeration tank, the SRT can be controlled by wasting a given percentage of the aeration tank volume each day.


19

2.6.4 Mixed liquor solids concentration and solids production The solids production from a biological reactor represents the mass of material that must be removed each day to maintain the process. It is of interest to quantify the solids production in terms of TSS, VSS, and biomass. By definition, the SRT also provides a convenient expression to calculate the total sludge produced daily from the activated-sludge process:

SRT

VXP T

VSSXT=, (2-41)

Where PXT, VSS = total solids wastes daily, g VSS/d XT = total MLVSS concentration in aeration tank, g VSS/m3 V = volume of rector, m3

SRT = solid retention time, d Because the 1/SRT in Eq. (2-41) represents the fraction of solids wasted per day and the mixed liquor can be assumed to be a homogeneous mixture of biomass and other solids, Eq. (2-41) can be used to calculate the amount of solids wasted for any of the mixed liquor components. For the amount of biomass wasted per day (Px), the biomass concentration X can be used in place of XT in Eq. (2-41). Mixed liquor solids concentration The total MLVSS in the aeration tank equals the biomass concentration X plus the nbVSS concentration Xi: XT = X + Xi (2-42) A mass balance is needed to determine the nbVSS concentration in addition to the active biomass VSS concentration. The MLVSS nbVSS concentration is affected by the amount of nbVSS in the influent wastewater, the amount of nbVSS wasted per day, and the amount of cell debris produced from cell decay. A materials balance on the inert material is as follows: Accumulation = inflow – outflow + generation

VrSRT

VXQXV

dt

dXix

iio

i,, +−=⎟

⎠⎞

⎜⎝⎛ (2-43)

Where Xo,i = nbVSS concentration in influent, g/m3 Xi = nbVSS concentration in aeration tank, g/m3 rX,i = rate of nbVSS production from cell debris, g/m3.d At steady-state (dXi/dt = 0) and substituting Eq. (2-21) for rX,i in Eq. (2-43) yields:

VrSRT

VXQX iX

iio ,,0 +−= (2-44)

SRTXkfSRT

XX ddioi ×+×=τ, (2-45)

Combining Eq. (2-39) and Eq. (2-45) for X and Xi produce the following equation that can be used to determine the total MLVSS concentration:


20

( )

ττSRTX

SRTXkfSRTk

SSYSRTX io

ddd

oT

×+×+

×+−××= ,

1 (2-46)

Heterotrophic Cell Non-biodegradable biomass debris VSS in influent

Solids production By substituting Eq. (2-46) for XT in Eq. (2-41) and replacing τ with V/Q, the amount of VSS produced and wasted daily is determined as follows:

( )iodd

d

oVSSX QXXVkf

SRTk

SSQYP ,, 1

++×+−= (2-47)

Fig. 2.2 Biodegradable soluble COD, biomass, and MLVSS concentrations versus SRT for complete-mix activated-sludge process (Metcalf and Eddy, 2003).

Eq. (2-39) is substituted for the biomass concentration (X) in Eq. (2-47) to show the daily VSS production rate in terms of the substrate removed, influent nbVSS, and kinetic coefficient as follows:

( ) ( )io

d

odd

d

oVSSX QX

SRTk

SRTSSYQkf

SRTk

SSQYP ,, 11

++

−++

−= (2-48)

A B C Heterotrophic Cell Non-biodegradable

biomass debris VSS in influent The effect of SRT on the performance of an activated-sludge system for soluble substrate removal is illustrated on Fig. 2.2. In addition to the soluble substrate concentration, the total VSS concentration which includes nabs is also shown. As the SRT increases, more biomass decays and thus more cell debris accumulates, so that the difference between MLVSS and biomass VSS concentration increases with SRT. Also illustrated on Fig. 2.2 is the fact that the soluble substrate concentration is very low (< 5 mg/L) at SRTs above 2 d. The low substrate concentration is typical of the activated-sludge process when used for the treatment of municipal wastewaters, and illustrates how effectively the organic compounds are degraded in the activated-sludge process.


21

The total mass of dry solids wasted per day is based on the TSS, which includes the VSS plus inorganic solids. Inorganic solids are in the influent wastewater (TSS-VSS) and the biomass contains 10 to 15 percent inorganic solids by dry weight. The influent inorganic solids are not soluble, and are assumed captured in the mixed liquor solids and removed in the wasted solids. Eq. (2-48) is modified to calculate the solids production in terms of TSS by adding the influent inorganic solids and by calculating the biomass in terms of TSS by assuming a typical biomass VSS/TSS ratio of 0.85. The ratio of VSS/TSS may vary from 0.80 to 0.90.

( )ooTSSX VSSTSSQCBA

P −+++=85.085.0, (2-49)

Where PX,TSS = net waste activated sludge produced each day, measured in terms of total

suspended solids, kg/d TSSo = influent wastewater TSS concentration, g/m3 VSSo = influent wastewater VSS concentration, g/m3 The mass of MLVSS and MLSS can be ontained by using Eqs. (2-48) and (2-49), respectively, with Eq. (2-41) as follows: Mass of MLVSS = XVSS.V = PX,VSS.SRT (2-50) Mass of MLSS = XTSS.V = PX,TSS.SRT (2-51) By selecting an appropriate MLSS concentration, the aeration volume can be determined from Eq. (2-51). MLSS concentration in the range of 2000 to 4000 mg/L may be selected and must be compatible with the sludge settling characteristics and clarifier design. 2.6.5 The observed yield The observed yield Yobs is based on the amount of solids production measured relative to the substrate removal, and may be calculated in terms of g TSS/g bsCOD or g BOD, or relative to VSS as g VSS/g bsCOD or gBOD. The measured solids production is the sum of the solids in the system effluent flow and the solids intentionally wasted, which equals the term Px defined in Eqs. (2-41), (2-48) and (2-49). The observed yield for VSS can be calculated by dividing Eq. (2-48) by the substrate removal rate, which is Q(So-S):

SS

X

SRTk

SRTYkf

SRTk

YY

o

io

d

dd

dobs −

+×+×+

×+= ,

11 (2-52)

Where Yobs = g VSS/g substrate removed For wastewaters with no nbVSS in the influent the solids production consists of only active biomass and cell debris, and the observed yield for VSS is as follows:

SRTk

SRTYkf

SRTk

YY

d

dd

dobs ×+

×+×+

=11

(2-53)

Heterotrophic biomass

Cell debris

Non-biodegradable VSS in influent


22

The impact of non-biodegradable influent VSS in Eq. (2-52) on the observed yield depends on the wastewater characteristics and the type of pretreatment. The effluent substrate concentration is generally very low compared to So and the term Xo,i/(So –S) can be approximated by Xo,i/So values range from 0.10 to 0.30 g/g with primary treatment and 0.30 to 0.50 without primary treatment. 2.6.6 Oxygen requirement The oxygen required for the biodegradation of carbonaceous material is determined from a mass balance using the bCOD concentration of the wastewater treated and the amount of biomass wasted from the system per day. If all of the bCOD were oxidized to CO2 and H2O, the oxygen demand would equal the bCOD concentration, but bacteria only oxidize a portion of the bCOD to provide energy and use a portion of the bCOD for cell growth. Oxygen is also consumed for endogenous respiration, and the amount will depend on the system SRT. For a given SRT, a mass balance on the system can be done where the bCOD removal equals the oxygen used plus the biomass VSS remaining (in terms of an oxygen equivalent), as was shown by Eq. (2-7). Thus, for a suspended growth process, the oxygen used is: Oxygen used = bCOD removed – COD of waste sludge (2-54) Ro = Q(So – S) – 1.42 Px,bio (2-55) Where Ro = oxygen required, kg/d Px,bio = biomass as VSS wasted per day, kg/d It is important to note that Px,bio includes active biomass and cell debris derived from cell growth and is thus the sum of terms A and B in Eq. (2-48). 2.6.7 Design and operating parameters In the mass balance for the complete-mix reactor, presented above, the SRT was introduced as the fundamental process parameter that affects the treatment efficient and general performance for the activated-sludge process. Two other activated-sludge process parameters used for the design and operation of the activated-sludge process, process parameters used for the design and operation of the activated-sludge process, the food to microorganism ratio and the volumetric loading rate, are introduce below. Food to microorganism (F/M) ratio The F/M ratio is defined as the rate of BOD or COD applied per unit volume of mixed liquor: Total applied substrate rate QSo F/M = ----------------------------------- = ------ (2-56) Total microbial biomass VX and So F/M = ---- (2-57) τX Where F/M = food to biomass ratio, g BOD or bsCOD/g VSS.d Q = influent wastewater flow-rate, m3/d


23

So = influent BOD or bsCOD concentration, g/m3 V = aeration tank volume, m3

X = mixed liquor biomass concentration in the aeration tank, g/m3 τ = hydraulic retention time of aeration tank, V/Q, d

Specific substrate utilization rate The F/M ratio can be related to the specific substrate utilization rate U, defined earlier, by the process efficiency: (F/M)E U = --------- (2-58) 100 Where E = BOD or bsCOD process removal efficiency as defined by Eq. (2-59): So - S E, % = --------- x 100 (2-59) So Substituting Eq. (2-56) for the F/M ratio and the term [(So-S)/So]100 for the process efficiency yields the specific substrate utilization rate U, as given previously: So - S U = ------- τX The value of U can also be calculated by dividing –rsu in Eq. (2-8) by X: kS U = --------- (2-61) Ks + S By combining Eqs. (2-61) and (2-35): 1 ----- = YU - kd (2-62) SRT Again, the net specific growth rate µ is also equal to 1/SRT. In turn, as given by Eq. (2-62), the specific growth rate is equal t the synthesis yield times the specific substrate utilization rate minus the endogenous decay rate. By substituting Eq. (2-58) into Eq. (2-62) the SRT can be related to the F/M ratio as follows: 1 E ----- = Y(F/M) ------ - kd (2-63) SRT 100 For systems designed for the treatment of municipal wastewater with activated-sludge SRT values in the 20- to 30- d range, the F/M value may range from 0.10 to 0.05 g BOD/g VSS.d, respectively. At SRT in the range of 5 to 7 d, the value may range from 0.3 to 0.5 g BOD/g VSS.d, respectively.,


24

Organic volumetric loading rate The organic volumetric loading rate, defined as the amount of BOD or COD applied to the aeration tank volume per day, is:

kggV

QSL o

org /103×= (2-64)

Where Lorg = volumetric organic loading, kg BOD/m3.d Q = influent wastewater flow-rate, m3/d So = influent BOD concentration, g/m3 V = aeration tank volume, m3 2.6.8 Process performance and stability The effects of the kinetics considered above on the performance and stability of the system shown on Fig. 2.3 will now be examined further. It was shown in Eqs. (2-61) and (2-33) that 1/SRT, the net microorganism specific growth rate, and U, the specific substrate utilization ratio, are related directly. For a specified waste, a given biological community, and a particular set of environmental conditions, the kinetic coefficients Y, k, Ks and kd, are fixed. It is important to note that domestic wastewater may have significant variability in its composition and may not always be treated as a single waste type in evaluating the kinetic coefficients. For given values of the coefficients, the effluent substrate concentration from the reactor is a direct function of the SRT, as shown on Eq (2-36). Setting the SRT value fixes the values of U and µ and also defines the efficiency of biological waste stabilization. Equation (2-36) for substrate is plotted on Fig. 2.3 for a growth-specified complete-mix system with recycle. As shown, the treatment efficiency and the substrate concentration are related directly to the SRT, and the reactor hydraulics (i.e., complete-mix or plus-flow).

Fig. 2.3 Effluent substrate concentration and removal efficiency for complete-mix and plus-flow reactors with recycle versus SRT (Metcalf and Eddy, 2003).

It can also be seen from Fig. 2.3 that there is a certain value of SRT below which waste stabilization does not occur. The critical SRT value is called the minimum solids retention residence time SRTmin. Physically, SRTmin is the residence time at which the cells are washed out or wasted from the system faster than they can reproduce. The minimum SRT can be calculated using Eq. (2-35), in which S = So. When washout occurs, the influent concentration So is equal to the effluent waste concentration S.


25

dos

o kSK

YkS

SRT−

+=

min

1 (2-65)

In many situations encountered in waste treatment, So is much greater than Ks so that Eq. (2-64) can be rewritten to yield:

dkYkSRT

−~1min

(2-66) or dm kSRT

−µ~1min

(2-67)

Equations (2-66) and (2-67) can be used to determine the SRTmin. Biological treatment processes should not be designed with SRT values equal to SRTmin. To ensure adequate waste treatment, biological treatment processes are usually designed and operated with a design SRT value from 2 to 20 times SRTmin. In effect, the ratio of the design SRT (SRTdes) to SRT min can be considered to be a process safety factor SF against system failure.

minSRT

SRTSF des= (2-68)

2.7 AEROBIC BIOLOGICAL OXIDATION 2.7.1 Process description The removal of BOD can be accomplished in a number of aerobic suspended growth or attached growth treatment processes. Both require sufficient contact time between the wastewater and heterotrophic microorganisms, and sufficient oxygen and nutrients. During the initial biological uptake of the organic material, more than half of it is oxidized and the remainder is assimilated as new biomass, which may by further oxidized by endogenous respiration. For both suspended and attached growth processes, the excess biomass produced each day is removed and processed to maintain proper operation and performance. The biomass is separated from the treated effluent by gravity separation, and more recent designs using membrane separation are finding applications. 2.7.2 Microbiology A wide variety of microorganisms are found in aerobic suspended and attached growth treatment processes used for the removal of organic material. Aerobic heterotrophic bacteria found in these processes are able to produce extra-cellular biopolymers that result in the formation of biological flocs (or biofilms for attached growth processes) that can be separated from the treated liquid by gravity settling with relatively low concentrations of free bacteria and suspended solids. Protozoa also play an important role in aerobic biological treatment processes. By consuming free bacteria and colloidal particulates, protozoa aid effluent clarification. Protozoa require a longer SRT than aerobic heterotrophic bacteria, prefer dissolved oxygen concentration above 1.0 mg/L, and are sensitive to toxic materials. Thus, their presence is a good indicator of a trouble-free stable process operation. Because of their size, protozoa can easily be observed with a light microscope at 10 to 200 magnifications. Rotifers can also be found in activated sludge and in biofilms, as well as nematodes and other multi-cellular microorganism. These organisms occur at longer biomass retention times, and their importance has not been well defined.


26

Aerobic attached growth processes depending on the biofilm thickness, generally have a much more complex microbial ecology that activated sludge with films containing bacteria, fungi, protozoan, rotifers, and possibly annelid worms, flatworms, and nematodes (WEF, 2000). Depending on process loadings and environmental conditions, a number of nuisance organisms can also develop in the activated sludge process. The principal problem caused by nuisance organism is a condition known as bulking sludge, in which the biological floc has poor settling characteristics. In the extreme, bulking sludge can result in high effluent suspended solids concentrations and poor treatment performance. Another nuisance condition, foaming, has been related to the development of two bacteria genera Nocardia and Microthrix (Pitt and Jenkins, 1990), which have hydrophobic cell surfaces and attach to air bubble surfaces, where they stabilize the bubbles to cause foam (see Fig. 2.4). The organisms can be found at high concentrations in the foam above the activated-sludge liquid.

(a) (b) Fig. 2.4 Examples of foam caused by Nocardia accumulated on the surface of activated sludge aeration

tanks (Metcalf and Eddy, 2003). 2.7.3 Stoichometry of aerobic biological oxidation The stoichiometry for aerobic oxidation was discussed previously but is repeated here for completeness. In aerobic oxidation, the conversion of organic mater is carried out by mixed bacterial cultures in general accordance with the stoichiometry shown below. Oxidation and synthesis: Bacteria COHNS + O2 + nutrients CO2 + NH3 + C5H7NO2 + other end products (2-69) Organic new cell Matter Endogenous respiration: Bacteria C5H7NO2 + 5O2 5CO2 + 2H2O + NH3 + energy (2-70) 113 160 1 1.42 I Eq. (2-69), COHNS is used to represent the organic matter in wastewater, which serve as the electron donor while the oxygen serves as the electron acceptors. Although the endogenous respiration reaction (2-70) is shown as resulting in relatively simple end products and energy, stable organic end products are also formed. From Eq. (2-5), it was shown that if all of the cells


27

(i. e. the electron donor) were oxidized completely, the UBOD or COD of the cells is equal to 1.42 times the concentration of cells as VSS. At longer SRT values, a greater portion of the cells will be oxidized. 2.7.4 Growth kinetics The form of the rate expressions for substrate utilization and biomass growth for the heterotrophic oxidation of organic substrates, based on the stoichiometry given above, were presented previously but are repeated below for ease of reference.

SK

kXSr

ssu +

−=

XkSK

kSXYXkYrr d

sdsug −

+=−−=

Both of the above expressions are of the saturation type. As noted previously, these expressions are similar to the saturation equation proposed by Monod (1942) for growth ad the Michaelis-Menten equation for substrate utilization (Bailey and Ollis, 1986). Typical k and Ks values at 20oC vary from 8.0 to 12.0 g COD/g VSS.d and 10 to 40 g bsCOD/m3, respectively. The Ks value can vary depending on the nature and complexity of the bsCOD components. For easily biodegradable single substrates, Ks values of less than 1.0 mg bsCOD/L have been measured (Bielefeldt and Stensel, 1999). Applying the above expressions for substrate utilization and biomass growth leads to the development of a series of design parameters including the solids retention time (SRT), the food to microorganism ratio (F/M), and the specific utilization rate (U). These design parameters are applied to the design of a variety of activated-sludge processes. With the exception of some difficult-to-degrade constituents in industrial wastewaters, the kinetics for aerobic oxidation of organic substrates seldom control the SRT design value for the activated-sludge process. For good floc formation, sufficient time is needed for the biomass in the activated-sludge aeration tank to develop extracellular polymers and a floc structure. More optimal flocculation and TSS removal in clarification occur typically at SRT values greater than 2.5 to 3.0 d at 20oC and 3 to 5 d at 10oC. However, some wastewater-treatment plants in warmer climates operate at SRT values varying from less than 1.0 to 1.5 d. Excessively long SRTs (> 20 d( may lead to floc deterioration with the development of small pinpoint floc particles that produce a more turbid effluent. However, even with pinpoint floc, effluent suspended solids concentrations of less than 30 g/m3 are generally achieved. The SRT may be varied in treatment plant operations to find the most optimal settling condition. 2.7.5 Environmental factors For carbonaceous removal, pH in the range of 6.0 to 9.0 is tolerate, while optimal performance occurs near a neutral pH. A reactor DO concentration of 2.0 mg/L is commonly used, and at concentrations above 0.50 mg/L there is little effect of the DO concentration on the degradation rate. For industrial wastewaters care must be taken to assure that sufficient nutrients (N and P) are available for the amount of bsCOD to be treated. Heterotrophic bacteria responsible for BOD removal can tolerate higher concentrations of toxic substances as compared to the bacteria responsible for ammonia oxidation or the production of methane.


28

2.8 BIOLOGICAL NITRIFICATION Nitrification is the term used to describe the two-step biological process in which ammonia (NH4-N) is oxidized to nitrite (NO2-N) and nitrite is oxidized to nitrate (NO3-N). The need for nitrification in wastewater treatment arises from water quality concerns over: (1) the effect of ammonia on receiving water with respect to DO concentrations and fish toxicity, (2) the need to provide nitrogen removal to control eutrophication, and (3) the need to provide nitrogen control for water-reuse applications including groundwater recharge. For reference, the current (2001) drinking water maximum contaminant level (MCL) for nitrate nitrogen is 45 mg/L as nitrate or 10 mg/L as nitrogen. The total concentration of organic and ammonia nitrogen concentration in municipal wastewaters is typically in the range from 25 to 45 mg/L as nitrogen based on flow-rate of 450 L/capita.d (120 gal/capita.d). In many parts of the world with limited water supplies, total nitrogen concentrations in excess of 200 mg/L as N have been measured in domestic wastewater. 2.8.1 Process description As with BOD removal, nitrification can be accomplished in both suspended growth and attached growth biological processes. For suspended growth processes, a more common approach is to achieve nitrification along with BOD removal in the same single-sludge process, consisting of an aeration tank, clarifier and sludge recycle system (see Fig. 2.5a). In cases where there is a significant potential for toxic and inhibitory substances in the wastewater, a two-sludge suspended growth system may be considered (see Fig. 2.5b). The two-sludge system consists of two aeration tanks and two clarifiers in series with the first aeration tank/clarifier unit operated at a short SRT for BOD removal. The BOD and toxic substance are removed in the first unit, so that nitrification can proceed unhindered in the second. A portion of influent wastewater usually has efficient solid flocculation and clarification. Because the bacteria responsible for nitrification grow much more slowly that heterotrophic bacteria, systems designed for nitrification generally have much longer hydraulic and solid retention times than those for systems deigned only for BOD removal. In attached growth systems used for nitrification, most of the BOD must be removed before nitrifying organisms can be established. The heterotrophic bacteria have a higher biomass yield and thus can dominate the surface are of fixed-film systems over nitrifying bacteria. Nitrification is accomplished in an attached growth reactor after BOD removal or in a separate attached growth system designed specifically for nitrification. 2.8.2 Microbiology Aerobic autotrophic bacteria are responsible for nitrification in activated sludge and biofilm processes. Nitrification, as noted above, is a two-step process involving two groups of bacteria. In the first state, ammonia is oxidized to nitrate by one group of autotrophic bacteria. In the second state, nitrite is oxidized to nitrate by another group of autotrophic bacteria. It should be noted that the two groups of autotrophic bacteria are distinctly different. Starting with classical experiments on nitrification by Winogradsky (1981), the bacteria genera commonly noted for nitrification in wastewater treatment are the autotrophic bacteria Nitrosomonass and Nitrobacter, which oxidize ammonia to nitrate and then to nitrate, respectively. Other autotrophic bateria genera capable of obtaining energy from the oxidation of ammonia to nitrite (prefix with Nitroso-) are Nitrosococcus, Nitrispira, Nitrosolobus, and Nitrisorobrio (Painter, 1970). It should be noted that during the 1990s, many more autotrophic bacteria were identified as being capable of oxidizing ammonia.


29

Besides Nitrobacter, nitrite can also be oxidized by other autotrophic bacteria (Nitro-) genera: Nitricoccus, Nitrospira, Nitrospina and Nitroeystis. Using oligo-nucleotide probes for ammonia-oxidizing bacteria, Wagner et al. (1995) showed that Nitrosomonas was common in activated sludge systems. For nitrite oxidation in activated sludge, Teske et al. (1994) found that Nitrococcus was quite prevalent. Whether different growth conditions can select for different genera of nitrifying bacteria or if heir nitrification kinetics are significantly different is unknown at present.

Fig. 2.5 Process configuration used for biological nitrification: (a) single-sludge suspended growth

system and (b) two-sludge suspended growth system (Metcalf and Eddy, 2003). 2.8.3 Stoichiometry of biological nitrification The energy-yielding two-step oxidation of ammonia to nitrate is as follows: Nitroso-bacteria: 2NH4

+ + 3O2 2NO2- + 4H+ + 2H2O (2-71)

Nitro-bacteria: 2NO2

- + O2 2NO3- (2-72)

Total oxidation reaction: NH4

+ + 2O2 NO2 + 2H+ + H2O (2-73) Based on the above total oxidation reaction, the oxygen required for complete oxidation of ammonia is 4.57 g O2/g N oxidized with 3.43 g O2/g used for nitrate production and 1.14 g O2/g NO2 oxidized. When synthesis is considered, the amount of oxygen required is less than 4.57 g O2/ g N. In addition to oxidation, oxygen is obtained from fixation of carbon dioxide and nitrogen into cell mass.


30

Neglecting cell tissue, the amount of alkalinity required to carry out the reaction given in Eq. (2-73) can be estimated by writing Eq. (2-73) as follows: NH4

+ + 2HCO3- + 2O2 NO3

- + 2CO2 + 3H2O (2-74) In the above equation, for each g of ammonia nitrogen (as N) converted, 7.14 g of alkalinity as CaCO3 will be required [2 x 50 g CaCO3/eq)/14]. Along with obtaining energy, a portion of the ammonium ion is assimilated into cell tissue. The biomass synthesis reaction can be represented as follows: 4CO2 + HCO3

- + NH4+ + H2O --> C5H7O2N + 5O2 (2-75)

As noted previously, the chemical formula C5H7O2N is used to represent the synthesized bacteria cells. The wastewater nitrogen concentration, BOD concentration, alkalinity, temperature, and potential for toxic compounds are major issues in the design of biological nitrification processes. Nitrifying bacteria need CO2 and phosphorus for cell growth, as well as trace elements. With such a low cell yield, the CO2 in air is adequate and phosphorus is seldom limiting. Trace element concentrations that have been found to stimulate the growth of nitrifying bacteria in pure culture work are Ca = 0.5, Cu = 0.01, Mg = 0.03, Mo = 0.001, Ni = 0.10 and Zn = 1.0 mg/L (Poduska, 1973). 2.8.4 Growth kinetics For nitrification systems operated at temperatures below 28oC, ammonia-oxidation kinetics versus nitrite-oxidation kinetics are rate-limiting, so that designs are based on saturation versus nitrite-oxidation s given below, assuming excess DO is available:

dnn

nmn k

NK

N −⎟⎟⎠

⎞⎜⎜⎝

⎛+

= µµ (2-76)

Where µn = specific growth rate of nitrifying bacteria, g new cells/g cells.d µnm = maximum specific growth rate of nitrifying bacteria, g new cells/g cells.d

N = nitrogen concentration, g/m3 kn = half-velocity constant, substrate concentration at one-half the maximum specific substrate utilization rate, g/m3 kdn = endogenous decay coefficient for nitrifying organisms, g VSS/g VSS.d

A wide range of maximum specific growth rates have been reported as a function of temperature (Randall et al., 1992). At 20oC, reported µnm varies from 0.25 to 0.77 g VSS/ g VSS/ The wide range of nitrification growth rates may be due to the presence of inhibitory substances in the wastewater and/or variations in experimental techniques and methods of analysis. In any event, the µnm values for nitrifying organisms are much lower than the corresponding values for heterotrophic organisms, requiring much longer SRT values for nitrifying activated – sludge systems. Typical design SRT values may range from 10 to 20 d at 10oC to 4-7 d at 20oC. Above 28oC, both ammonia and nitrite oxidation kinetics should be considered. At elevated temperatures, the relative kinetics of NH4-N and NO2

--N oxidation change, and NO2-N will accumulate at lower SRT values.


31

For fully acclimated complete-mix activated sludge nitrification systems, at temperatures below 25oC with sufficient Do present, the NO2

--N concentration may be less than 0.10 mg/L as compared to NH4-N concentrations in the range of 0.50 to 1.0 mg/L. However, during the initiation of nitrification, NO2-N concentrations will be greater than NH4-N concentrations, as the growth of nitrite-oxidizing bacteria cannot occur until the ammonia-oxidizing bacteria generate nitrite. Under transient conditions, NO2-N concentrations of 5 to 20 mg/L are possible. Nitrification rates are affected by the liquid DO concentration in activated sludge. In contrast to what has been observed for aerobic heterotrophic bacteria degradation of organic compounds, nitrification rates increase up to DO concentrations of 3 to 4 mg/L. To account for the effects of DO, the expression for the specific growth rate (Eq. 2-76) is modified as follows:

mon

nmn k

DOK

DO

NK

N −⎟⎟⎠

⎞⎜⎜⎝

⎛+⎟⎟

⎠

⎞⎜⎜⎝

⎛+

= µµ (2-77)

Where DO = dissolved oxygen concentration, g/m3 K0 = half-saturation coefficient for DO, g/m3 Other terms as defined previously. While the above kinetic model and coefficients, based on observed results, can be used to describe nitrification in system at low modest organic loadings, the use of these models will generally overpredict nitrification rates in system with high organic loading rate. Stenstrom and Song (1991) have shown experimentally that the effect of DO on nitrification is affected by the activated-sludge floc size and density, and total oxygen demand of the mixed liquor. Nitrifying bacteria are distributed within a floc containing heterotrophic bacteria and other solids, with floc diameter ranging from 100 to 400 µm. Oxygen from the bulk liquid diffused in to floc particles and bacteria deeper within the floc are exposed to lower DO concentration. At higher organic loadings, there is a greater substrate in the mixed liquor, which cause a high oxygen consumption rate within the floc. Therefore, a higher bulk liquid DO concentration is need maintain the same internal floc DO concentration and subsequence nitrification rates. At low DO concentration (<0.50 mg/L) where nitrification rate are greatly inhibited, the low DO inhibition effect has been shown to be greater for Nitrobacter than for Nitrosomonas. In such case, incomplete nitrification will occur with increase NO2-N concentration in the effluent. The presence of nitrite in the effluent is particularly troublesome for plants that use chlorination for disinfection, as nitrite is readily oxidized by chlorine requiring 4g chlorine/g NO2-N. Environment factors Nitrification is affected by a number of environmental factor including pH, toxicity, metals, and unionized ammonia. Hydrogen-Ion Concentration (pH). Nitrification is pH sensitive and rates decline significantly at pH values below 6.8. At pH values near 5.8 to 6.0, the rates may be 10 to 20 percent of the rate at pH 7.0 (U.S. EPA, 1993). Optimal nitrification rates occur at pH values in the 7.5 to 8.0 range. A pH of 7.0 to 7.2 is normally used to maintain reasonable nitrification rate, and for locations with low-alkalinity water, alkalinity is added at the wastewater-treatment plant to maintain acceptable pH values. The amount of alkalinity added depend on the initial alkalinity concentration and the amount of NH4-N to be oxidized. Alkalinity may be added in the form of lime, soda ash, sodium bicarbonate, or magnesium hydroxide depending on cost and chemical handling issues.


32

Toxicity. Nitrifying organism are sensitive to wide range of organic and inorganic compounds and at concentration well below those concentrations that would affect aerobic heterotrophic organisms. In many case, nitrification rate are inhibited even thought bacteria continue to grow and oxidize ammonia and nitrite, but at significantly reduced rates. In some case, toxicity may be sufficient to kill the nitrifying bacteria. Nitrifiers have been shown to be good indicators of the presence of organic toxic compounds at low concentration (Blum and Speece, 1991). Toxic organic compounds have been listed by Hockenbury and Grady (1977) and Sharma and Ahlert (1977). Compounds that are toxic include solvent organic chemicals, amines, proteins, tannins, phenolic compounds, alcohols, cyanates, ethers, carbamates, and benzene. Because of the numerous compounds that can inhibit nitrification, it is difficult to pinpoint the source of nitrification toxicity for wastewater plants whit inhibition, and extensive sampling of the collection system is normally needed to find the source. Metals. Metals are also of concern for nitrifiers and Skinner and Walker (1961) have shown complete inhibition of ammonia oxidation at 0.25 mg/L nickel, 0.25 mg/L chromium, and 0.10 mg/L copper. Un-ionized ammonia. Nitrification is also inhibited by un-ionized ammonia (NH3) or free ammonia, and un-ionized nitrous acid (HNO2). The inhibition effects are dependent on the total nitrogen species concentration, temperature, and pH. At 20oC and pH 7.0, the NH4 concentrations at 100 mg/L and 20 mg/L may initiate inhibition of NH4-N and NO2-N oxidation, respectively, and NO2-N concentrations at 280 mg/L may initiate inhibition of NO2-N oxidation (US.EPA, 1993). 2.9 BIOLOGICAL DENITRIFICATION The biological reduction of nitrate to nitric oxide, nitrous oxide, and nitrogen gas is termed denitrification. Biological denitrification is an integral part of biological nitrogen removal, which involves both nitrification and denitrification. Compared to alternatives of ammonia stripping, breakpoint chlorination, and ion exchange, biological nitrogen removal is generally more cost-effective and used more often. Biological nitrogen removal is used in wastewater treatment where there are concerns for eutrophication and where ground water must be protected against elevated NO3

—N concentrations where wastewater treatment plant effluent used for groundwater recharge and other reclaimed water applications. 2.9.1 Process description Two modes of nitrogen removal can occur in biological processes, and these are termed assimilating and dissimilating nitrate reduction (see Fig. 2.6). Assimilating nitrate reduction involves the reduction of nitrate to ammonia for use in cell synthesis. Assimilation occurs when NH4-N is not available and is independent of DO concentration. On the other hand, dissimilating nitrate reduction or biological denitrification is coupled to the respiratory electron transport chain, and nitrate and nitrite is used as an electron acceptor for the oxidation of a variety of organic or inorganic electron donors. Two basic flow diagrams for activated-sludge denitrification and the conditions that drive the denitrification reaction rates are illustrated in Fig. 2.7. The first flow diagram (Fig. 2.7a) is for the Modified Ludzak-Ettinger (MLE) process (U.S. EPA, 1993), the most common process used for biological nitrogen removal in municipal wastewater treatment. The process consists of an anoxic tank followed by the aeration tank where nitrification occurs. Nitrate produced in the


33

aeration tank is recycled back to the anoxic tank. Because the organic substrate in the influent wastewater provides the electron donor for oxidation reduction reactions using nitrate, the process is termed substrate denitrification. Further, because the anoxic process precedes the aeration tank, the process is known as a preanoxic denitrification.

Fig. 2.6 Nitrogen transformations in biological treatment processes (Metcalf and Eddy, 2003).

Fig. 2.7 Type of denitrification processes and the reactors used for their implementation: (a) substrate

driven (preanoxic denitrification) and (b) endogenous driven (postanoxic denitrification) (Metcalf and Eddy, 2003).

In the second process shown on Fig. 2.7b, denitrification occurs after nitrification and the electron donor source is from endogenous decay. The process illustrated on Fig. 2.7b is generally termed a postanoxic denitrification as BOD removal has occurred first and is not available to drive the nitrate reduction reaction. When a postanoxic denitrification process depends solely on endogenous respiration for energy, it has much slower rate of reaction than for the preanoxic processes using wastewater BOD. Often an exogenous carbon source such as methanol or acetate is added to postanoxic processes to provide sufficient BOD for nitrate reduction and to increase the rate of denitrification. Postanixic processes include both suspended and attached growth


34

granular-medium filtration process, both nitrate reduction and effluent suspended solid removal occurs in the same reactor. The denitrification preanoxic and postanoxic processes described employ heterotrophic bacteria for nitrate reduction, but other pathways for biological nitrogen removal exist. Ammonia can be converted to nitrogen gas by novel autotrophic bacteria under anaerobic conditions and by heterotrophic-nitrifying bacteria under aerobic conditions. The organisms identified with these reactions are presented in the Microbiology section below. Littleton et al. (2000) looked for possible influence of the autotrophic denitrifying bacteria or to heterotrophic nitrifying bacteria for nitrogen removal in aerated full-scale, long SRT, single-sludge, activated-sludge systems but could not find any significant activity for them. However, the use of novel autotrophic denitrifying bacteria has been demonstrated in a process for the treatment of high ammonia strength aerobic digestion centrate in a fluidized bed reactor at 30 to 35oC (Strous et al. 1997, Jetten et al. 1999). The fluidized bed reactor is effective in maintaining these relatively slow growing bacteria. Under anaerobic conditions NH4

+ is oxidized by NO2- to produce nitrogen gas

and small amount of nitrate. About 1.3 moles of NO2- are used per mole of NH4

+ (Strous et al. 1999a). A portion of the centrate stream must be nitrified to nitrite, but no carbon source is needed for nitrogen removal. The process is termed the Annamox process, which stands for anaerobic ammonium oxidation. The oxidation of ammonia to nitrite in anaerobic digester centrate has been done in a process termed the SHARON process, which can be used with the Annamox process (Jetten et al. 1999). 2.9.2 Microbiology A wide range of bacteria has been shown capable of denitrification, but similar microbial capability has not been found in algae or fungi. Bacteria capable of denittrification are both autotrophic and heterotrophic. The heterotrophic organisms include the following genera: Achromobacter, Acinetobacter, Agrobacterium, Alcaligenes, Arthrobacter, Bacillus, Chromobacterium, Corynebacterium, Flavobacterium, Hypomicrobium, Moraxella, Neisseria, Paracoccus, Propionibacterium, Pseudomonas, Rhizobium,Rhodopseudomonas, Spirillum, and Vibrio (Payne, 1981). In addition, Gayle (1989) list Halobacterium and Methanomonas, Pseudomonas species are the most common and widely distributed of all the denitrifiers, and have been shown to use a wide array of organic compounds including hydrogen, methanol, carbonhydrates, organic acids, alcohols, benzoates, and other aromatic compounds (Payne, 1981). Most of these bacteria are facultative aerobic organisms with the ability to use oxygen as well as nitrate or nitrite, and some can also carry out fermentation in the absence of nitrate or oxygen. Other autotrophic bacteria that can denitrify use hydrogen and reduced sulfur compounds as electron donors during denitrification. Both groups of organisms can grow heterotrophically if an organic carbon source is present (Gayle, 1989). Nitrogen removal can also be accomplished by the heterotrophic and autotrophic-nitrifying bacteria under certain conditions, and by a unique bacteria associated with the Annamox process, which was discovered in the mid-1990s. Denitrification can occur under anaerobic condition by heterotrophic nitrifying bacteria (Robertson and Kuenen, 1990, and Patureau et al. 1994), so that simultaneous nitrification and denitrification exist with the conversion of ammonia to gaseous nitrogen products. The heterotrophic bacteria, Paracocus pantotropha, have been studied extensively for simultaneous oxidation and nitrate reduction. The oxidation of ammonia by heterotrophic bacteria requires energy, which can be obtained by nitrate or nitrite reduction by P.pantotropha under aerobic conditions. A readily available substrate, such as acetate, is also needed. Due to the need for the carbon substrate, which is in limited supply in aerobic activated-sludge systems, little growth of heterotrophic nitrifiers is expected (van Loosdrech and Jetten, 1998) in aerobic wastewater treatment systems.


35

Autotrophic nitrifying bacteria, such as Nitrosomonas europaea, can use nitrite to oxidize ammonia, with production of nitrogen gas, when dissolved oxygen is not present (Boc et al. 1995). With oxygen present, these bacteria oxidize the ammonia with oxygen as electron acceptors. Ammonia oxidation with the reduction of nitrite under anaerobic conditions has also been shown at temperatures above 20oC in the Annamox process (Strous et al. 1997). The bacteria in Annamox process are different than the autotrophic nitrifying bacteria described above, in that it cannot use oxygen for ammonia oxidation (Jetten et al. 1999). The Annamox bacteria could not be isolated and grow in pure culture (Strous et al. 1999b), but an enrichment was obtained by density purification for 16S rRNA extraction and analysis. Phylogenetic analysis showed that the bacteria is in the order Planctomycetales, a division with domain Bacteria. Under anaerobic conditions the ammonia oxidation rate by the Annamox bacteria was shown to be 6 to 10 times faster than that for N. Europaea (Jetten et al. 1999). 2.9.3 Stoichiometry of Biological Denitrification Biological denitrification involves the biological oxidation of many organic substrates in wastewater treatment using nitrate or nitrite as the electron acceptors instead of oxygen. In the absent of DO or under limited DO concentrations, the nitrate reductase enzyme in the electron transport respiratory chain is induced, and helps to transfer hydrogen and electrons to nitrate as the terminal electron acceptor. The nitrate reduction reactions involves the following reduction steps from nitrate to nitrite, to nitric oxide, to nitrous oxide, and to nitrogen gas: NO3

- NO2- NO N2O N2 (2-78)

In biological nitrogen removal process, the electron donor is typically one of three sources: (1) the bsCOD in the influent wastewater, (2) the bsCOD produced during endogenous decay, and (3) an exogenous source such as methanol or acetate. The latter has been added in separate treatment units, such as polishing filters, after nitrification where almost no bsCOD remains. The term C10H19O3N is often used to represent the biodegradable organic matter in wastewater (U.S. EPA, 1993). Wastewater: C10H19O3N + 10NO3

- 5N2 + 10CO2 + 3H2O + NH3 + 10OH- (2-79) Methanol: 5CH3OH + 6 NO3

- 3N2 + 5 CO2 + 7 H2O + 6 OH- (2-80) Acetate: 5 CH3COOH + 8 NO3

- 4N2 + 10 CO2 + 6 H2O +8 OH- (2-81) In all the above heterotrophic denitrification reactions, one equivalent of alkalinity is produced per equivalent of NO3-N reduced, which equates to 3.57g of alkalinity (as CaCO3) production per g of nitrate nitrogen reduced. Recall from nitrification that 7.14 alkalinity (as CaCO3) was consumed per g of NH4-N oxidized, so that by denitrification about one-half of the amount destroyed by nitrification can be recovered.


36

2.9.4 Growth kinetics For biological denitrification, the biokinetic equations used to describe bacteria growth and substrate utilization are similar to those described previously in this chapter for aerobic heterotrophic bacteria. The rate of soluble substrate utilization is also controlled by the soluble substrate concentration, with nitrate serving as the electron acceptor in lieu of oxygen. The nitrate concentration controls the substrate utilization kinetics only at very low NOrN concentrations, near 0.1 mg/L. Nitrate serves as an electron acceptor in the same way as oxygen from a biokinetics perspective, and thus the nitrate utilization rate (denitrification rate) is proportional to the substrate utilization rate. There are two general cases where the substrate utilization rate controls the denitrification rate. The first is for anoxic/aerobic processes where the organic substrate (electron donor) is from the influent wastewater fed into the anoxic reactor. The second is for postanoxic denitrification, where nitrate reduction is done after secondary treatment in a reactor receiving another carbon source. Because the BOD is depleted from secondary treatment, an exogenous carbon source is used to drive the nitrate reduction reaction. In most cases methanol is supplied to create the demand for the electron acceptor in suspended growth or attached growth processes. Endogenous respiration creates a demand for nitrate in addition to that caused by substrate utilization and oxidation. This reaction occurs in the mixed liquor of anoxic tanks and is at a much lower rate than the denitrification rate caused by substrate utilization. In the case where wastewater provides the electron donor, an anoxic reactor receives the influent wastewater and it is followed in the treatment system by an aerobic reactor where nitrification occurs. Heterotrophic bacterial growth occurs in both the anoxic and aerobic zones with nitrate and oxygen consumption, respectively. The mixed liquor biomass concentration can be calculated based on the total amount of BOD removed, but only a portion of that biomass can use both nitrate and oxygen as electron acceptors. The microorganisms in the other portion are strict aerobes and can only use oxygen as the electron acceptor. To apply biokinetic expressions for denitrification, the substrate utilization rate expression is modified to account for the fact that only a portion of the biomass is active in the anoxic zone. The substrate utilization rate rsu expression is modified by a term to show a lower utilization rate in the anoxic zone as follows:

SK

kXSr

Ssu +

−= η (2-82)

Where η = fraction of denitrifying bacteria in the biomass, g VSS/gVSS K, X, S, Ks are as defined previously. When nitrate is used as an electron acceptor instead of oxygen, the maximum specific substrate utilization rate (k) may be lower than the rate with oxygen as the electron acceptor. In Eq. (2-82), the k value used is the same as that used for oxygen, and any effect of lower kinetic rate is incorporated into the η term. The Ks value has been found to be similar, whether nitrate or oxygen is the electron acceptor (Stensel and Home, 2000). The value for η has been found to vary from 0.20 to 0.80 for preanoxic denitrification reactors fed domestic wastewaters (Stensel and Home, 2000). The activated sludge configuration, the system SRT, and the fraction of influent BOD removed with nitrate appear to affect the η value. For anoxic/aerobic processes with substantial substrate and nitrate removal in the preanoxic zone η may be close to 0.80. In spite of the fact that only a portion of the mixed liquor biomass can


37

use nitrate, the anoxic reactor volumes used for anoxic/aerobic processes range from only 10 to 30 percent of the total volume (anoxic plus aerobic) for treating domestic wastewater. For postanoxic suspended growth or attached growth processes the biomass is developed under mainly anoxic conditions and with a selected single organic substrate. In this case the η term is not necessary because the biomass consists of mainly denitrifying bacteria. The biokinetic equations presented previously can then be used with the appropriate kinetic coefficient values (k, Ks, Y, kd) to design a postanoxic complete-mix suspended growth process. The kinetic coefficient values for growth using methanol have been developed at 10°C and 20°C in laboratory studies (Randall et aI., 1992). The kinetics for methanol utilization are such that the SRTs required for a denitrification suspended growth process are in the same range as SRTs for aerobic systems designed for BOD removal only, about 3 to 6 d. Effect of dissolved oxygen concentration Dissolved oxygen can inhibit nitrate reduction by repressing the nitrate reduction enzyme. In activated-sludge flocs and biofilms, denitrification can proceed in the presence of low bulk liquid DO concentrations. A dissolved oxygen concentration of 0.2 mg/L and above has been reported to inhibit denitrification for a Pseudomonas culture (Skerman and MacRae, 1957; Terai and Mori, 1975) and by Dawson and Murphy (1972) for activated-sludge treating domestic wastewater. Nelson and Knowles (1978) reported that denitrification ceased in a highly dispersed growth at a DO concentration of 0.13 mg/L. The effect of nitrate and DO concentration on the biokinetics is accounted for by two correction factors to Eq. (2-82) expressed in the form of two saturation terms as follows:

( )η⎟⎟⎠

⎞⎜⎜⎝

⎛+⎟

⎟⎠

⎞⎜⎜⎝

⎛

+⎟⎟⎠

⎞⎜⎜⎝

⎛+

−=DOK

K

NOK

NO

SK

kXSr

o

o

NOSSsu '

'

3,

3

3

(2-83)

Where Ko

’ = DO inhibition coefficient for nitrate reduction, mg/L KS,NO3 = half velocity coefficient for nitrate limited reaction, mg/L Other terms are as defined previously. The value of Ko

’ is system-specific. Values in the range from 0.1 to 0.2 mg/L have been proposed for Ko

’ and 0.1 mg/L for Ks.NO3 (Barker and Dold, 1997). Assuming a Ko value of 0.1 mg/L the rate of substrate utilization with nitrate as the electron acceptor at DO concentrations of 0.10, 0.20, and 0.50 mg/L would be at 50, 33, and 17 percent of the maximum rate, respectively. Effect of simultaneous nitrification-denitrification In activated-sludge systems, the issue of DO concentration is confounded by the fact that the measured bulk liquid DO concentration does not represent the actual DO concentration within the activated-sludge floc. Under low DO concentration conditions, denitrification can occur in the floc interior, while nitrification is occurring at the floc exterior. Also in activated sludge tanks operated at low DO concentrations, both aerobic and anaerobic zones exist depending on mixing conditions and distance from the aeration point, so that nitrification and denitrification can be occurring in the same tank. Under these conditions, nitrogen removal that occurs in a single aeration tank is referred to as simultaneous nitrification and denitrification. Although both nitrification and denitrification are occurring at reduced rates as indicated by the DO effects described for both processes, if a sufficient SRT and HRT exist, the overall nitrogen removal can be significant. Rittman and Langeland (1985) reported greater than 90 percent nitrogen removal by nitrification and denitrification in an activated-sludge system used to treat municipal


38

wastewater at DO concentrations below 0.50 mg/L and with values of HRT greater than 25 hours. Environmental factors Alkalinity is produced in denitrification reactions and the pH is generally elevated, instead of being depressed as in nitrification reactions. In contrast to nitrifying organisms, there has been less concern about pH influences on denitrification rates. No significant effect on the denitrification rate has been reported for pH between 7.0 and 8.0, while Dawson and Murphy (1972) showed a decrease in the denitrification rate as the pH was decreased from 7.0 to 6.0 in batch un acclimated tests. 2.10 BIOLOGICAL PHOSPHORUS REMOVAL The removal of phosphorus by biological means is known as biological phosphorus removal. Phosphorus removal is generally done to control eutrophication because phosphorus is a limiting nutrient in most freshwater systems. Treatment plant effluent discharge limits have ranged from 0.10 to 2.0 mg/L of phosphorus depending on plant location and potential impact on receiving waters. Chemical treatment using alum or iron salts is the most commonly used technology for phosphorus removal, but since the early 1980s success in full-scale plant biological phosphorus removal has encouraged further use of the technology. The principal advantages of biological phosphorus removal are reduced chemical costs and less sludge production as compared to chemical precipitation. 2.10.1 Process description In the biological removal of phosphorus, the phosphorus in the influent wastewater is incorporated into cell biomass, which subsequently is removed from the process as a result of sludge wasting. Phosphorus accumulating organisms (PAOs) are encouraged to grow and consume phosphorus in systems that use a reactor configuration that provides PAOs with a competitive advantage over other bacteria. The reactor configuration utilized for phosphorus removal is comprised of an anaerobic tank having a HRT value of 0.50 to 1.0 h that is placed ahead of the activated-sludge aeration tank (see Fig. 2.8). The contents of the anaerobic tank are mixed to provide contact with the return activated sludge and influent wastewater. Anaerobic contact tanks have been placed in front of many different types of suspended growth processes, with aerobic SRT values ranging from 2 to 40 d. Phosphorus removal in biological systems is based on the following observations (Sedlak, 1991): 1. Numerous bacteria are capable of storing excess amounts of phosphorus as polyphosphates in

their cells. 2. Under anaerobic conditions, PAOs will assimilate fermentation products (e.g., volatile fatty

acids) into storage products within the cells with the concomitant release of phosphorus from stored polyphosphates.

3. Under aerobic conditions, energy is produced by the oxidation of storage products and polyphosphate storage within the cell increases.


39

Fig. 2.8 Biological phosphorus removal: (a) typical reactor configuration. Photos below flow diagram are

of (b) transmission electron microscope images of polyhydroxybutyrate and (c) polyphosphate storage granules (Metcalf and Eddy, 2003).

A simplified version of the processes occurring in the anaerobic and aerobic/anoxic reactors or zones is presented below. In many applications for phosphorus removal, an anoxic reactor follows the anaerobic reactor and precedes the aerobic reactor. Most PAOs can use nitrite in place of oxygen to oxidize their stored carbon source. A more comprehensive description of the biochemistry and intracellular transformations can be found in Wentzel et al. (1991). Processes occurring in the anaerobic zone - Acetate is produced by fermentation of bsCOD which, as defined earlier, is dissolved

degradable organic material that can be assimilated easily by the biomass. Depending on the value of T for the anaerobic zone, some colloidal and particulate COD is also hydrolyzed and converted to acetate, but the amount is generally small compared to that from the bsCOD conversion.

- Using energy available from stored polyphosphates, the PAOs assimilate acetate and produce

intracellular polyhydroxybutyrate (PHB) storage products. Some glycogen contained in the cell is also used. Concurrent with the acetate uptake is the release of orthophosphate (O-P04), as well as magnesium, potassium, and calcium cations.

- The PHB content in the PAOs increases while the polyphosphate decreases. Processes occurring in the aerobic/anoxic zone - Stored PHB is metabolized, providing energy from oxidation and carbon for new cell growth. - Some glycogen is produced from PHB metabolism.


40

- The energy released from PHB oxidation is used to form polyphosphate bonds in cell storage so that soluble orthophosphate (O-P04) is removed from solution and incorporated into polyphosphates within the bacterial cell. Cell growth also occurs due to PHB utilization and the new biomass with high polyphosphate storage accounts for phosphorus removal.

- As a portion of the biomass is wasted, stored phosphorus is removed from the biotreatment reactor for ultimate disposal with the waste sludge.

The events occurring in the anaerobic and aerobic zones are illustrated graphically on Fig. 2.9.

Fig. 2.9 Fate of soluble BOD and phosphorus in nutrient removal reactor (adapted from Sedlak, 1991; Metcalf and Eddy, 2003).

2.10.2 Microbiology Phosphorus is important in cellular energy transfer mechanisms via adenosine triphosphate (ATP) and polyphosphates. As energy is produced in oxidation reduction reactions, adenosine diphosphate (ADP) is converted to ATP with 7.4 kcal/mole of energy captured in the phosphate bond. As the cell uses energy, ATP is converted to ADP with phosphorus release. For common heterotrophic bacteria in activated-sludge treatment the typical phosphorus composition is 1.5 to 2.0 percent. However, many bacteria are able to store phosphorus in their cells in the form of energy-rich polyphosphates, resulting in phosphorus content as high as 20 to 30 percent by dry weight. The polyphosphates are contained in volutin granules within the cell along with Mg2+, Ca2+, and K+ cations. In the anaerobic zone, concentrations of O-PO4 as high as 40 mg/L can be measured in the liquid, as compared to wastewater influent concentrations of 5 to 8 mg/L. The high concentration of O-PO4 can be taken as an indication that phosphorus release by the bacteria has occurred in this zone. Also in this zone, significant amounts of poly-bhydroxybutyrate (PHB) are found stored in bacteria cells, but the PHB concentration declines appreciably in the subsequent anoxic and/or aerobic zones and can be measured and quantified. The O-PO4 is taken up from solution in the aerobic and anoxic zones, generally leading to very low remaining concentrations. Based on investigations of biological phosphorus removal, it was found that acetate was essential to forming the PHB under anaerobic conditions, which provided a competitive advantage for the PAOs. The anaerobic zone in the anaerobic/aerobic treatment process is termed a "selector," because it provides conditions that favor the proliferation of the PAOs, by the fact that a portion of the influent bCOD is consumed by the PAOs instead of other heterotrophic bacteria. Because the PAOs prefer low-molecular-weight fermentation product substrates, the preferred food source would not be available without the anaerobic zone that provides for the fermentation of the influent bsCOD to acetate. Because of the polyphosphate storage ability, the PAOs have energy available to assimilate the acetate in the anaerobic zone. Other aerobic heterotrophic bacteria


41

have no such mechanism for acetate uptake, and they are starved while the PAOs assimilate COD in the anaerobic zone. It should also be noted that the PAOs form very dense, good settling floc in the activated sludge, which is an added benefit. In some facilities, the anaerobic/aerobic process sequence has been used because of the sludge settling benefits, even though biological phosphorus removal was not required. Care must be taken in the handling of the waste sludge from biological phosphorus removal systems. When the sludge is held under anaerobic conditions, phosphorus release will occur. Release of O-PO4 is possible even without acetate addition as the bacteria use the stored polyphosphate for an energy source. The release of O-PO4 can also occur after extended contact time in the anaerobic zone of the biological phosphorus treatment system. In that case the released phosphorus may not be taken up in the aerobic zone because the release was not associated with acetate uptake and PHB storage for later oxidation. The release of O-PO4 under these conditions is termed secondary release (Barnard, 1984), which can lead to a lower phosphorus removal efficiency for the biological process. 2.10.3 Stoichiometry of biological Phosphorus removal Based on the description of the phosphorus removal mechanism, acetate uptake in the anaerobic zone is critical in determining the amount of PAOs that can be produced and, thus, the amount of phosphorus that can be removed by this pathway. If significant amounts of dissolved oxygen or nitrate enter the anaerobic zone, the acetate can be depleted before it is taken up by the PAOs, and treatment performance will be hindered. Biological phosphorus removal is not used in systems that are designed with nitrification without including a means for denitrification to minimize the amount of nitrate in the return sludge flow to the anaerobic zone. The amount of phosphorus removed by biological storage can be estimated from the amount of bsCOD that is available in the wastewater influent as most of the bsCOD will be converted to acetate in the short anaerobic hydraulic detention time HRT. The following assumptions are used to evaluate the stoichiometry of biological phosphorus removal: (1) 1.06 g acetate/g bsCOD will be produced as most of the COD fermented will be converted to VFAs due to the low cell yield of the fermentation process, (2) a cell yield of 0.30 g VSS/g acetate, and (3) a cell phosphorus content 0.3 g Pig VSS. Using these assumptions, about 10 g of bsCOD will be required to remove I g of phosphorus by the biological storage mechanism. Other bCOD removal in the activated-sludge system will result in additional phosphorus removal by normal cell synthesis. Better performance for biological phosphorus removal systems is achieved when bsCOD or acetate is available at a steady rate. Periods of starvation or low bsCOD concentrations result in changes in the intracellular storage reserves of glycogen, PHB, and polyphosphates and rapidly lead to decreased phosphorus removal efficiency (Stephens and Stensel, 1998). 2.10.4 Growth kinetics Biological phosphorus growth kinetics are within the same order of magnitude of other heterotrophic bacteria. Mamais and Jenkins (1992) showed that biological phosphorus removal could be maintained in anaerobic/aerobic systems at SRTs greater than 2.5 d at 20°C. A maximum specific growth rate at 20°C is given as 0.95 g/god (Barker and Dold, 1997).


42

2.10.5 Environmental factors System performance is not affected by DO as long as the aerobic zone DO concentration is above 1.0 mg/L. At pH values below 6.5, phosphorus removal efficiency is greatly reduced (Sedlak, 1991). In biological phosphorus removal systems, sufficient cations associated with polyphosphate storage must also be available. The recommended molar ratios of Mg, K, and Ca to phosphorus are 0.71, 0.50, and 0.25, respectively (Wentzel et aI., 1989). Thus, for an influent soluble phosphorus concentration of 10 mg/L, 5.6, 6.3, and 3.2 mg/L of Mg, K, and Ca, respectively, would be required. The relative amounts of these cations associated with phosphate storage are 0.28, 0.26, and 0.09 mole/mole of phosphorus, respectively (Sedlak, 1991). Most municipal wastewaters have sufficient amounts of these inorganic elements, but care must be taken to assure sufficient amounts in industrial applications or laboratory experiments. 2.11 ANAEROBIC FERMENTATION AND OXIDATION Anaerobic fermentation and oxidation processes are used primarily for the treatment of waste sludge (see Fig. 2.10) and high-strength organic wastes. However, applications for dilute waste streams have also been demonstrated and are becoming more common. Anaerobic fermentation processes are advantageous because of the lower biomass yields and because energy, in the form of methane, can be recovered from the biological conversion of organic substrates. Although most fermentation processes are operated in the mesophilic temperature range (30 to 35°C), there is increased interest in thermophilic fermentation alone or before mesophilic fermentation. The latter is termed temperature phased anaerobic digestion (TPAD) and is typically designed with a sludge SRT of 3 to 7 d in the first thermophilic phase at 50 to 60°C and 7 to 15 d in the final mesophilic phase (Han and Dague, 1997). Thermophilic anaerobic digestion processes, are used to accomplish high pathogen kill to produce Class A biosolids, which can be used for unrestricted reuse applications. For treating high-strength industrial wastewaters, anaerobic treatment has been shown to provide a very cost-effective alternative to aerobic processes with savings in energy, nutrient addition, and reactor volume. Because the effluent quality is not as good as that obtained with aerobic treatment, anaerobic treatment is commonly used as a pretreatment step prior to discharge to a municipal collection system or is followed by an aerobic process.

Fig. 2.10 Views of anaerobic digesters: (a) Kuwait City, Kuwait and (b) egg-shaped digester at Okinawa, Japan (Metcalf and Eddy, 2003).


43

2.11.1 Process description Three basic steps are involved in the overall anaerobic oxidation of a waste: (1) hydrolysis, (2) fermentation (also known as acidogenesis), and (3) methanogenesis. The three steps are illustrated schematically on Fig. 2.11. The starting point on the schematic for a particular application depends on the nature of the waste to be processed.

Fig. 2.11 Anaerobic process schematic of hydrolysis, fermentation, and methanogenesis (Metcalf and Eddy, 2003).

Hydrolysis. The first step for most fermentation processes, in which particulate material is converted to soluble compounds that can then be hydrolyzed further to simple monomers that are used by bacteria that perform fermentation, is termed hydrolysis. For some industrial wastewaters, fermentation may be the first step in the anaerobic process. Fermentation. The second step is fermentation (also referred to as acidogenesis). In the fermentation process, amino acids, sugars, and some fatty acids are degraded further, as shown on Fig. 2.11. Organic substrates serve as both the electron donors and acceptors. The principal products of fermentation are acetate, hydrogen, CO2 and propionate and butyrate. The propionate and butyrate are fermented further to also produce hydrogen, CO2 and acetate. Thus, the final products of fermentation (acetate, hydrogen, and CO2) are the precursors of methane formation (methanogenesis). The free energy change associated with the conversion of propionate and butyrate to acetate and hydrogen requires that hydrogen be at low concentrations in the system (H2 < 10-4 atm), or the reaction will not proceed (McCarty and Smith, 1986). Methanogenesis. The third step, methanogenesis, is carried out by a group of organisms known collectively as methanogens. Two groups of methanogenic organisms are involved in methane production. One group, termed aceticlastic methanogens, split acetate into methane and carbon dioxide. The second group, termed hydrogen-utilizing methanogens, use hydrogen as the electron donor and CO2 as the electron acceptor to produce methane. Bacteria within anaerobic processes, termed acetogens, are also able to use CO2 to oxidize hydrogen and form acetic acid. However, the acetic acid will be converted to methane, so the impact of this reaction is minor.


44

As shown on Fig. 2.12, about 72 percent of the methane produced in anaerobic digestion is from acetate formation.

Fig. 2.12 Carbon and hydrogen flow in anaerobic digestion process (Metcalf and Eddy, 2003). 2.11.2 Microbiology The group of nonmethanogenic microorganisms responsible for hydrolysis and fermentation consists of facultative and obligate anaerobic bacteria. Organisms isolated from anaerobic digesters include Clostridium spp., Peptococcus anaerobus, Bifidobacterium spp., Desulphovibrio spp., Corynebacterium spp., Lactobacillus, Actinomyces, Staphylococcus, and Escherichia coli. Other physiological groups present include those producing proteolytic, lipolytic, ureolytic, or cellulytic enzymes. The microorganisms responsible for methane production, classified as archaea, are strict obligate anaerobes. Many of the methanogenic organisms identified in anaerobic digesters are similar to those found in the stomachs of ruminant animals and in organic sediments taken from lakes and rivers. The principal genera of microorganisms that have been identified at mesophilic conditions include the rods (Methanobacterium, Methanobacillus) and spheres (Methanococcus, Methanothrix, and Methanosarcina). Methanosarcina and Methanothrix (also termed Methanosaeta) are the only organisms able to use acetate to produce methane and carbon dioxide. The other organisms oxidize hydrogen with carbon dioxide as the electron acceptor to produce methane. The acetate-utilizing methanogens were also observed in thermophilic reactors (van Lier, 1996; Zinder and Koch, 1984; and Ahring, 1995). Some species of Methanosarcina were inhibited by temperature at 65°C, while others were not, but no inhibition of Methanothrix was shown. For hydrogen-utilizing methanogens at temperatures above 60°C, Methanobacterium was found to be very abundant. Syntrophic relationships in fermentation The methanogens and the acidogens form a syntrophic (mutually beneficial) relationship in which the methanogens convert fermentation end products such as hydrogen, formate, and acetate to methane and carbon dioxide. Because the methanogens are able to maintain an extremely low partial pressure of H2, the equilibrium of the fermentation reactions is shifted toward the formation of more oxidized end products (e.g., formate and acetate). The utilization of the hydrogen produced by the acidogens and other anaerobes by the methanogens is termed interspecies hydrogen transfer. In effect, the methanogenic organisms serve as a hydrogen sink that allows the fermentation reactions to proceed. If process upsets occur and the methanogenic organisms do not utilize the hydrogen produced fast enough, the propionate and butyrate


45

fermentation will be slowed with the accumulation of volatile fatty acids in the anaerobic reactor and a possible reduction in pH. Nuisance organisms Nuisance organisms in anaerobic operations are the sulfate-reducing bacteria, which can be a problem when the wastewater contains significant concentrations of sulfate. These organisms can reduce sulfate to sulfide, which can be toxic to methanogenic bacteria at high enough concentrations. Where high sulfide concentrations occur, one solution is to add iron at controlled amounts to form iron sulfide precipitate. Sulfate-reducing bacteria, obligate anaerobes of the domain Bacteria, are morphologically diverse, but share the common characteristic of being able to use sulfate as an electron acceptor and are divided into one of two groups depending on whether they produce fatty acids or use acetate. Group I sulfate reducers can use a diverse array of organic compounds as their electron donor, oxidizing them to acetate and reducing sulfate to sulfide. A common genus found in anaerobic biochemical operations is Desulfovibrio. Group II sulfate reducers oxidize fatty acids, particularly acetate, to carbon dioxide, while reducing sulfate to sulfide. A bacteria commonly found in this group is in the genus Desulfobacter. 2.11.3 Stoichiometry of anaerobic fermentation and oxidation A limited number of substrates are used by the methanogenic organisms and reactions defined as CO2 and methyl group type reactions are shown as follows (Madigan et al., 1997), involving the oxidation of hydrogen, formic acid, carbon monoxide, methanol, methylamine, and acetate, respectively. 4H2 + CO2 CH4 + 2H2O (2-84) 4HCOO- + 4H+ CH4 + 3CO2 + 2H2O (2-85) 4CO + 2H2O CH4 + 3CO2 (2-86) 4CH3OH 3CH4 + CO2 + 2H2O (2-87) 4(CH3)3N + H2O 9CH4 + 3CO2 + 6H2O + 4NH3 (2-88) CH3COOH CH4 + CO2 (2-89) In the reaction for the aceticlastic methanogens as given by Eq. (2-89), the acetate is cleaved to form methane and carbon dioxide. A COD balance can be used to account for the changes in COD during fermentation. Instead of oxygen accounting for the change in COD, the COD loss in the anaerobic reactor is accounted for by the methane production. By stoichiometry the COD equivalent of methane can be determined. The COD of methane is the amount of oxygen needed to oxidize methane to carbon dioxide and water. CH4 + 2O2 CO2 + 2H2O (2-90) From the above, the COD per mole of methane is 2(32 gO2/mole) = 64 g O2/mole CH4. The volume of methane per mole at standard conditions (O°C and 1 atm) is 22.414 L, so the CH4 equivalent of COD converted under anaerobic conditions is 22.414/64 = 0.35 L CH4/g COD. 2.11.4 Growth kinetics In anaerobic processes two rate-limiting concepts are important: (1) the hydrolysis conversion rate and (2) the soluble substrate utilization rate for fermentation and methanogenesis. The hydrolysis of colloidal and solid particles does not affect the process operation and stability but


46

does affect the total amount of solids converted. In anaerobic digestion processes used for municipal waste sludges, greater than 30 days detention time is needed to approach full conversion of solids. The soluble substrate utilization kinetics are of great concern to develop a stable anaerobic process. Because of the relatively low free energy change for anaerobic reactions, growth yield coefficients are considerably lower than the corresponding values for aerobic oxidation. TYpical synthesis yield and endogenous decay coefficients for fermentation and methanogenic anaerobic reactions are Y = 0.10 and 0.04 g VSS/g VSS and kd = 0.04 and 0.02 g VSS/g VSS.d, respectively. The process is more stable when the volatile fatty acid (VFA) concentrations approach a minimal level, which can be taken as an indication that a sufficient methanogenic population exists and sufficient time is available to minimize hydrogen and VFA concentrations. The rate-limiting step is the conversion of VFAs by the methanogenic organisms and not the fermentation of soluble substrates by the fermenting bacteria. Thus, the methanogenic growth kinetics are of most interest in anaerobic process designs. Appropriate system SRTs are selected based on kinetics and treatment goals. At 20,25, and 35°C, the washout or SRTmin values for methanogenesis are 7.8, 5.9, and 3.2 d, respectively (Lawrence and McCarty, 1970). Thus, with a factor of safety of 5, design SRT values would be about 40, 30, and 15 days, respectively, for a suspended growth process. Safety factors higher than 5 have been used to provide a more stable process (Parker and Owen, 1986). 2.11.5 Environmental factors Anaerobic processes are sensitive to pH and inhibitory substances. A pH value near neutral is preferred and below 6.8 the methanogenic activity is inhibited. Because of the high CO2 content in the gases developed in anaerobic processes (30 to 35 percent CO2)' a high alkalinity is needed to assure pH near neutrality. An alkalinity concentration in the range of 3000 to 5000 mg/L as CaCO3 is often found. For sludge digestion sufficient alkalinity is produced by the breakdown of protein and amino acids to produce NH3, which combines with CO2 and H2O to form alkalinity as NH4(HCO3) For industrial wastewater applications, especially for waste containing mainly carbohydrates, it is necessary to add alkalinity for pH control. Substances inhibitory to anaerobic processes are NH3, H2S, and various other inorganic and organic compounds. 2.12 BIOLOGICAL REMOVAL OF TOXIC AND RECALCITRANT ORGANIC

COMPOUNDS Most of the organic compounds in domestic wastewater and some in industrial wastewaters are of natural origin and can be degraded by common bacteria in aerobic or anaerobic processes. However, currently there are over 70,000 synthetic organic chemicals, termed xenobiotic compounds, in general use (Schwarzenbach et al., 1993). Unfortunately, some of these organic compounds pose unique problems in wastewater treatment, due to their resistance to biodegradation and potential toxicity to the environment and human health. Organic compounds that are difficult to treat in conventional biological treatment processes are termed refractory. In addition, there are naturally occurring substances, such as those found in petroleum products, that are of similar concern.


47

2.12.1 Development of biological treatment methods Since the early 1970s, information and knowledge related to the biodegradation of toxic and refractory compounds has increased significantly, based on work with specific industrial wastewaters (i.e., petrochemical, textile, pesticide, pulp and paper, and pharmaceutical industries). In addition, since the 1980s, significant progress has also been made on the biodegradation of organic substances found at hazardous waste sites. Work in both of these fields has expanded knowledge on the capabilities and limitations of biodegradation. With a few exceptions most organic compounds can be biodegraded eventually, but in some cases the rates may be slow, unique environmental conditions may be required (i.e., redox potential, pH, temperature), fungi may be needed instead of prokaryotes, or specific bacteria capable of degrading the xenobiotic compounds may be needed. For example, anaerobic degradation of polychlorinated biphenyls (PCB) occurred using bacteria seed from sediment in the Hudson River where PCB had accumulated over decades, but after 1.5 years of exposure in a laboratory anaerobic digester used to treat municipal wastewater plant sludge, bacteria could not be developed to degrade PCB (Ballapragada et al., 1998). Importance of specific microorganisms The ability to degrade toxic and recalcitrant compounds will depend primarily on the presence of appropriate microorganism(s) and acclimation time. In some cases, special seed sources are needed to provide the necessary microorganisms. Once the critical microorganism is present, long-term exposure to the organic compound may be needed to induce and sustain the enzymes and bacteria required for degradation. Acclimation times can vary from hours to weeks depending on the microorganism population and organic compound. Melcer et al. (1994) found that a period of 3 weeks was required before complete removal of dichlorobenzene (DCB) occurred in a municipal activated-sludge plant. They noted that intermittent addition of DCB resulted in much lower treatment efficiencies. Strand et al. (1999) showed that after 4 weeks of constant exposure to dinitrophenol in a laboratory activated-sludge process, seeded from a municipal wastewater plant, dinitrophenol degradation increased from 0 to 98 percent. When dinitrophenol was not added to the process, the ability to degrade dinitrophenol was eventually lost. Thus, it appears that a relatively constant supply of toxic and recalcitrant organic compounds can lead to better biodegradation performance than intermittent additions. Biodegradation pathways The three principal types of degradation pathways that have been observed are (l) the compound serves as a growth substrate; (2) the organic compound provides an electron acceptor; and (3) the organic compound is degraded by cometabolic degradation. In cometabolic degradation, the compound that is degraded is not part of the microorganism's metabolism. Degradation of the compound is brought about by a nonspecific enzyme and provides no benefit to the cell growth. Complete biodegradation of toxic and recalcitrant organic compounds to harmless end products such as CO2 and H20 or methane may not always occur, and instead biotransformation to a different organic compound is possible. Care must be taken to determine if the organic compound produced is innocuous, or just as harmful as or more harmful than the initial compound. 2.12.2 Anaerobic degradation Many toxic and recalcitrant organic compounds are degraded under anaerobic conditions, with the compound serving as a growth substrate with fermentation and ultimate methane production. Typical examples include nonhalogenated aromatic and aliphatic compounds such as phenol,


48

toluene, alcohols, and ketones. However, most chlorinated organic compounds are not attacked easily under anaerobic conditions and do not serve as growth substrates. Fortuitously, many of these compounds also serve as electron acceptors in anaerobic oxidation reduction reactions. Most of the work and application for anaerobic degradation of chlorinated organic compounds have been related to subsurface contamination of chlorinated solvents at hazardous waste sites (McCarty, 1999). Examples of chlorinated compounds degraded under anaerobic conditions include tetrachloroethene, trichloroethene, carbon tetrachloride, trichlorobenzene, pentachlorophenol, chlorohydrocarbons, and PCBs. The chlorinated compound serves as the electron acceptor, and hydrogen produced from fermentation reactions provides the main electron donor. Hydrogen replaces chlorine in the molecule, and such reactions have generally been referred to as anaerobic dehalogenation or anaerobic dechlorination. For example, dechlorination of tetrachloroethene proceeds sequentially with a loss of chlorine in each step via trichloroethene to dichloroethene to vinyl chloride and finally to ethene. As the number of chlorine molecules on the organic molecule decreases, the reactions tend to be slower and less complete. Dechlorination of tetrachloroethene, trichlorobenzene, and pentachlorophenol has been demonstrated in lab-scale anaerobic digesters (Ballapragada et aI., 1998) treating municipal primary and secondary sludge. However, the reaction rates were slow with mono- and dichlorophenol and mono-I dichlorobenzenes remaining. Conversion of tetrachloroethene to vinyl chloride and ethene occurred in the digesters after one year of acclimatization and constant exposure of the chloroethenes. 2.12.3 Aerobic biodegradation With proper environmental conditions, seed source, and acclimation time, a wide range of toxic and recalcitrant organic compounds have been found to serve as growth substrates for heterotrophic bacteria. Such compounds include phenol, benzene, toluene, poly aromatic hydrocarbons, pesticides, gasoline, alcohols, ketones, methylene chloride, vinyl chloride, munitions compounds, and chlorinated phenols. However, many chlorinated organic compounds cannot be attacked readily by aerobic heterotrophic bacteria and thus do not serve as growth substrates. Some of the lesser chlorinated compounds, such as dichloromethane, 1,2-dichloroethane, and vinyl chloride can be used as growth substrates by aerobic bacteria. Fortunately, a number of chlorinated organic compounds are degradable by cometabolic degradation. It should be noted that organic compounds that are saturated fully with chlorine are degraded only by anaerobic dechlorination (Stensel and Bielefeldt, 1997). Chlorinated organic compounds that have been degraded by cometabolic degradation include trichloroethene, dichloroethene, vinyl chloride, chloroform, dichloromethane, and trichloroethane. Cometabolic degradation is possible by bacteria that produce nonspecific mono-oxygenase or dioxygenase enzymes. These enzymes mediate a reaction with oxygen and hydrogen and change the structure of the chlorinated compound. Bacteria that produce oxygenase enzymes oxidize certain substrates that induce the enzyme. Oxygenase-producing bacteria include methanotrophic bacteria that oxidize methane, a number of bacteria that can oxidize phenol or toluene, a number of bacteria that can oxidize propane, and nitrifying bacteria that oxidize ammonia to nitrite. The reaction of the nonspecific oxygenase enzyme with the organic chlorinated compound typically produces an intermediate compound that is degraded by other aerobic heterotrophic


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bacteria in the biological consortia. Various reactor designs have been developed to apply this biological process for treatment of contaminated groundwater or vapor extraction gas streams (Lee et al., 2000). While such reactions are possible in municipal and industrial biological wastewater treatment processes, a large amount of the chlorinated organic compounds that may be present are more likely lost from the process by volatilization during aeration, because of their high volatility and the minimal potential for cometabolic bacteria to be present. 2.13 BIOLOGICAL REMOVAL OF HEAVY METALS Metal removal in biological treatment processes is mainly by adsorption and complexation of the metals with the microorganisms. In addition, processes that result in transformations and precipitation of metals are possible. Microorganisms combine with metals and adsorb them to cell surfaces because of interactions between the metal ions and the negatively charged microbial surfaces. Metals may also be complexed by carboxyl groups found in microbial polysaccharides and other polymers, or absorbed by protein materials in the biological cell. The removal of metals in biological processes has been found to fit adsorption characteristics displayed by the Freundlich isotherm model (Mullen et al., 1989; Kunz et al., 1976). A significant amount of soluble metal removal has been observed in biological processes, with removals ranging from 50 to 98 percent depending on the initial metal concentration, the biological reactor solids concentrations, and system SRT. In anaerobic processes the reduction of sulfate to hydrogen sulfide can promote the precipitation of metal sulfides. A classic example is the addition of ferric or ferrous chloride to anaerobic digesters to remove sulfide toxicity by the formation of iron sulfide precipitates.

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CHAPTER 3

FUNDAMENTALS OF BIOLOGICAL SOLID WASTE TREATMENT TECHNOLOGY

3.1 OVERVIEW OF BIOLOGICAL SOLID WASTE TREATMENT Before considering the specific processes employed for the biological conversion of wastes, it will be helpful to review (1) the general nutritional requirements of the microorganisms commonly encountered in solid waste conversion facilities, (2) the type of microbial metabolism based on the need for molecular oxygen, (3) the types of microorganisms of importance in the conversion of solid waste, (4) environmental requirements, (5) aerobic and anaerobic transformations, and (6) process selection. 3.1.1 Nutritional requirements for microbial growth To continue to reproduce and function properly, an organism must have a sources of energy; carbon for the synthesis of new cell tissue, and inorganic elements (nutrients) such as nitrogen, phosphorus, sulfur, potassium, calcium, and magnesium. Organic nutrients (growth factors) may also be required for cell synthesis. Carbon and energy sources, usually referred to as substrates, and nutrient and growth factor requirements for various types of organisms are considered in the following discussion. Nutrient and growth factor requirements Nutrients, rather than carbon or an energy source, may at times be the limiting material for microbial cell synthesis and growth. The principal inorganic nutrients needed by microorganisms are nitrogen, sulfur, phosphorus, potassium, magnesium, calcium, iron, sodium, and chlorine. Minor nutrients of importance include zinc, manganese, molybdenum, selenium, cobalt, copper, nickel, and tungsten. In addition to the inorganic nutrients just cited, some organisms may also need organic nutrients. Required organic nutrients, known as growth factors, are compounds needed by an organism as precursors or constituents of organic cell material that cannot be synthesized from other carbon sources. Although growth factor requirements differ from one organism to another, the major growth factors fall into the following three classes: (1) amino acids, (2) purines and pyrimidines, and (3) vitamins. Microbial nutrition and biological conversion processes The major objective in most biological conversion processes is the conversion of the organic matter in the waste to a stable end product. In accomplishing this type of treatment, the chemoheterotrophic organisms are of primary importance because of their requirement for organic compounds as both carbon and energy source. The organic fraction of municipal solid waste typically contains adequate amounts of nutrients (both inorganic and organic) to support the biological conversion of the waste. With some commercial wastes, however, nutrients may not be present in sufficient quantities. In these cases, nutrient addition is necessary for the proper bacterial growth and for the subsequent degradation of the organic waste.

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3.1.2 Types of microbial metabolism Chemoheterotrophic organisms may be further grouped according to their metabolic type and their requirement for molecular oxygen. Organisms that generate energy by enzyme-mediated electron transport from an electron donor to an external electron acceptor (such as oxygen) are said to have a respiratory metabolism. In contrast, fermentative metabolism does not involve the participation of an external electron acceptor. Fermentation is a less efficient energy-yielding process than respiration; as a consequence, heterotrophic organisms that are strictly fermentative are characterized by lower growth rates and cell yields than respiratory heterotrophs. When molecular oxygen is used as the electron acceptor in respiratory metabolism, the process is known as aerobic respiration. Organisms that are dependent on aerobic respiration to meet their energetic needs can exist only when there is a supply of molecular oxygen. These organisms are called obligate aerobic. Oxidized inorganic compounds such as nitrate and sulfate can function as electron acceptors for some respiratory organisms in the absence of molecular oxygen. In environmental engineering, processes that make use of these organisms are often referred to as anoxic. Organisms that generate energy by fermentation and that can exist only in an environment that is devoid of oxygen are obligate anaerobic. There is another group of microorganisms, which has the ability to grow in either the presence or the absence of molecular oxygen. These organisms are called facultative anaerobes. The facultative organisms fall into two subgroups, based on their metabolic abilities. True facultative anaerobes can shift from fermentative to aerobic respiratory metabolism, depending upon the presence or absence of molecular oxygen. Aero-tolerant anaerobes have a strictly fermentative metabolism but are relatively insensitive to the presence of molecular oxygen. 3.1.3 Types of microorganisms Microorganisms are commonly classified, on the basis of cell structure and function, as eukaryotes, eubacteria, and archaebacteria. The prokaryotic groups (eubacteria and archaebacteria) are of primary importance in biological conversion of the organic faction of solid wastes and are generally referred to simply as bacteria. Bacteria Bacteria are single cells-spheres, rods, or spirals. Spherical forms (cocci) vary from 0.5 to 4 µm in diameter; rods (bacilli) are from 0.5 to 20 µm long and 0.5 to 4 µm wide; spirals (spirilla) may be more than 10 p.m long and about 0.5 µm wide. Bacteria are ubiquitous in nature and are found in aerobic (in the presence of oxygen) and anaerobic (in the absence of oxygen) environments. Because of the wide variety of inorganic and organic compounds that can be used by bacteria to sustain growth, bacteria are used extensively in a variety of industrial operations to accumulate intermediate and end products of metabolism. Tests on a number of different bacterial species indicate that they are about 80 percent water and 20 percent dry material, of which 90 percent is organic and 10 percent is inorganic. An approximate empirical formula for the organic fraction is C5H7NO2. On the basis of this formula, about 53 percent by weight of the organic fraction is carbon. Compounds that make up the inorganic portion include P2O5 (50 percent), CaO (9 percent), Na2O (11 percent), MgO (8 percent), K2O (6 percent), and Fe2O3 (1 percent). Because all these elements and compounds must be derived from the environment, a shortage of these substances would limit, and in some cases alter, the growth of bacteria.

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Fungi Fungi are considered to be multicellular, nonphotosynthetic, heterotrophic protists. Most fungi have the ability to grow under low-moisture conditions, which do not favor the growth of bacteria. In addition, fungi can tolerate relatively low pH values. The optimum pH value for most fungal species appears to be about 5.6, but the viable range is from 2 to 9. The metabolism of these organisms is essentially aerobic, and they grow in long filaments, called hyphae, composed of nucleated cell units and varying in width from 4 to 20 µm. Because of their ability to degrade a wide variety of organic compounds over a broad range of environmental conditions, fungi have been used extensively in industry for the production of valuable compounds, such as organic acids (e.g., citric, gluconic), various antibiotics (e.g., penicillin, griseofulvin), and enzymes (e.g., cellulase, protease, amylase). Yeasts Yeasts are fungi that cannot form filaments (mycelium) and are therefore unicellular. Some yeasts form elliptical cells 8 to 15 µm by 3 to 5 µm, whereas others are spherical, varying in size from 8 to 12 µm in diameter. In terms of industrial processing operations, yeasts may be classified as "wild" and "cultured." In general, wild yeasts are of little value, but cultured yeasts are used extensively to ferment sugars to alcohol and carbon dioxide. Actinomycetes The actinomycetes are a group of organisms with intermediate properties between bacteria and fungi. They are similar in form to fungi, except that the width of the cell is only 0.5- to 1.4 µm. In industry, this group of microorganisms is used extensively for the production of antibiotics. Because their growth characteristics are similar, actinomycetes are often grouped with fungi for discussion purposes. 3.1.4 Environmental requirements Environmental conditions of temperature and pH have an important effect on the survival and growth of microorganisms. In general, optimal growth occurs within a fairly narrow range of temperature and pH values, although the microorganism may be able to survive within much broader limits. For instance, temperatures below the optimum typically have a more significant effect on the bacterial growth rate than temperatures above the optimum. It has been observed that growth rates double with approximately every woe increase in temperature until the optimum temperature is reached. According to the temperature range in which they function best, bacteria may be classified as psychrophilic, mesophilic, or thermophilic. The hydrogen ion concentration, expressed as pH, is not a significant factor in the growth of microorganisms, in and of itself, within the range from 6 to 9 (which represents a thousandfold difference in the hydrogen ion concentration). Generally, the optimum pH for bacterial growth lies between 6.5 and 7.5. However, when the pH goes above 9.0 or below 4.5, it appears that the undissociated molecules of weak acids or bases can enter the cell more easily than hydrogen and hydroxide ions and, by altering the internal pH, damage the cell. Moisture content is another essential environmental requirement for the growth of microorganisms. The moisture content of the organic wastes to be converted must be known, especially if a dry process such as composting is to be used. In many composting operations, it has been necessary to add water to obtain optimum bacterial activity. The addition of water in

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anaerobic fermentation processes will depend on the characteristics of the organic waste and the type of anaerobic process that is used. The biological conversion of an organic waste requires the biological system to be in a state of dynamic equilibrium. To establish and maintain dynamic equilibrium, the environment must be free of inhibitory concentrations of heavy metals, ammonia, sulfides, and other toxic constituents. 3.1.5 Aerobic biological transformations The general aerobic transfonnation of solid waste can be described by means of the following equation: Bacteria Organic matter + O2 + nutrients new cells + resistant organic matter + CO2 + H2O + NH3 + SO4

2- + … + heat If the organic matter in solid waste is represented (on a molar basis) as CaHbOcNd, the production of new cells and sulfate is not considered, and the composition of the resistant material is represented (on a molar basis) as CwHxOyNz, then the amount of oxygen required for the aerobic stabilization of the biodegradable organic fraction of MSW can be estimated by using the following equation: CaHbOcNd + 0.5 (ny + 2s + r – c)O2 nCwHxOyNz + sCO2 + rH2O + (d-nx)NH3 The tenns CaHbOcNd and CwHxOyNz represent the empirical mole composition of the organic material initially present and at the conclusion of the process. If complete conversion is accomplished, the corresponding expression is 4a + b – 2c – 3d b - 3d CaHbOcNd + ---------------------- O2 aCO2 + -------- +H2O + dNH3 4 2 In many cases the ammonia, NH3, produced from the carbonaceous oxidation of organic matter is oxidized further to nitrate, NO3 - (a process known as nitrification). The amount of oxygen required for the oxidation of ammonia to nitrate can be computed by the following equations: NH3 + 3/2O2 HNO2 + H2O HNO2 + 1/2O2 JNO3 ------------------------------------ NH3 + 2O2 H2O + HNO3 3.1.6 Anaerobic biological transformations The production of methane from solid wastes by anaerobic digestion, or anaerobic fermentation as it is often called, is described in the following discussion. Process microbiology. The biological conversion of the organic fraction of municipal solid waste under anaerobic conditions is thought to occur in three steps. The first step in the process involves the enzyme-mediated transfonnation (hydrolysis) of higher-molecular-mass compounds into compounds suitable for use as a source of energy and cell tissue. The second step involves the bacterial conversion of the compounds resulting from the first step into identifiable lower-

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molecular-mass intennediate compounds. The third step involves the bacterial conversion of the intennediate compounds into simpler end products, principally methane and carbon dioxide. In the anaerobic decomposition of wastes, a number of anaerobic organisms work together to bring about the conversion of organic portion of the wastes to a stable end product. One group of organisms is responsible for hydrolyzing organic polymers and lipids to basic structural building blocks such as fatty acids, monosaccharides, amino acids, and related compounds. A second group of anaerobic bacteria fennents the breakdown products from the first group to simple organic acids, the most common of which in anaerobic digestion is acetic acid. This second group of microorganisms, described as nonmethanogenic, consists of facultative and obligate anaerobic bacteria that are often identified in the literature as "acidogens" or "acid fonners." A third group of microorganisms converts the hydrogen and acetic acid fonned by the acid fonners to methane gas and carbon dioxide. The bacteria responsible for this conversion are strict anaerobes, called methanogenic, and are identified in the literature as "methanogens" or "methane fonners." Many of the methanogenic organisms identified in landfills and anaerobic digesters are similar to those found in the stomachs of ruminant animals and in organic sediments taken from lakes and river. The most important bacteria of the methanogenic group are the ones that utilize hydrogen and acetic acid. They have very slow growth rates; as a result, their metabolism is usually considered rate-limiting in the anaerobic treatment of an organic waste. Waste stabilization in anaerobic digestion is accomplished when methane and carbon dioxide are produced. Methane gas is highly insoluble, and its departure from a landfill or solution represents actual waste stabilization. Biochemical pathways. It is important to note that methane bacteria can only use a limited number of substrates for the formation of methane. Currently, it is known that methanogens use the following substrates: CO2 + H2, formate, acetate, methanol, methylamines, and carbon monoxide. Jypical energy-yielding conversion reactions involving these compounds are as follows: 4H2 + CO2 CH4 + 2H2O 4HCOOH CH4 + 3CO2 + 2H2O CH3COOH CH4 + CO2 4CH3OH 3CH4 + CO2 + 2H2O 4(CH3)3N + 6H2O 9CH4 + 3CO2 + 4NH3 4CO + 2H2O CH4 + 3CO2 In anaerobic fermentation, the two principal pathways involved in the formation of methane are (1) the conversion of carbon dioxide and hydrogen to methane and water and (2) the conversion of formate and acetate to methane, carbon dioxide, and water. The methanogens and the acidogens form a syntrophic (mutually beneficial) relationship in which the methanogens convert fermentation end products such as hydrogen, formate, and acetate to methane and carbon dioxide. The methanogens are able to utilize the hydrogen produced by the acidogens, because of their efficient hydrogenase. Because the methanogens are able to maintain an extremely low partial pressure of H2, the equilibrium of the fermentation reactions is shifted towards the formation of more oxidized end products (e.g., formate and acetate). The utilization of the hydrogen, produced by the acidogens and other anaerobes, by the methanogens is termed interspecies hydrogen transfer. In effect, the methanogenic bacteria remove compounds that would inhibit the growth of acidogens. Environmental factors. To maintain an anaerobic treatment system that will stabilize an organic waste efficiently, the nonmethanogenic and methanogenic bacteria must be in a state of dynamic equilibrium. To establish and maintain such a state, the reactor contents should be void of

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dissolved oxygen and free from inhibitory concentrations of free ammonia and such constituents as heavy metals and sulfides. Also, the pH of the aqueous environment should range from 6.5 to 7.5. Sufficient alkalinity should be present to ensure that the pH will not drop below 6.2, because the methane bacteria cannot function below this point. When digestion is proceeding satisfactorily, the alkalinity will normally range from 1000 to 5000 mg/L and the volatile fatty acids will be less than 250 mg/L. Values for alkalinity and volatile fatty acids in the high-solids anaerobic digestion process can be as high as 12,000 and 700 mg/L, respectively. A sufficient amount of nutrients, such as nitrogen and phosphorus, must also be available to ensure the proper growth of the biological community. Depending on the nature of the sludges or waste to be digested, growth factors may also be required. Temperature is another important environmental parameter. The optimum temperature ranges are the mesophilic, 30 to 38°C (85 to 100°F), and the thermophilic, 55 to 60°C (131 to 140°F). Gas production. The general anaerobic transformation of solid waste can be described by means of the following equation. Organic matter + H2O + nutrients new cells + resistant organic matter + CO2 + CH4 + NH3 + H2S + heat For practical purposes, the overall conversion of the organic fraction of solid waste to methane, carbon dioxide, and ammonia can be represented with the following equation: CaHbOcNd nCwHxOyNz + mCH4 + sCO2 + rH2O + (d – nx)NH3 The terms CaHbOcNd and CwHxOyNz are used to represent (on a molar basis) the composition of the organic material present at the start and the end of the process, respectively. In operations where solid wastes have been mixed with wastewater sludge, it has been found that the gas collected from the digesters contains between 50 and 60 percent methane. It has also been found that about 10 to 16 ft3 of gas is produced per pound of biodegradable volatile solids destroyed. 3.1.7 Biological process selection Aerobic and anaerobic processes both have a place in solid waste management. Each process offers different advantages. In general, the operation of anaerobic processes is more complex than that of aerobic processes. However, anaerobic processes offer the benefit of energy recovery in the form of methane gas and thus are net energy producers. Aerobic processes, on the other hand, are net energy users because oxygen must be supplied for waste conversion, but they offer the advantage of relatively simple operation and, if properly operated, can significantly reduce the volume of the organic portion of MSW. The operational characteristics of both aerobic and anaerobic solid waste-processing systems are described in the following sections. 3.2 Aerobic composting Aerobic composting is the most commonly used biological process for the conversion of the organic portion of municipal solid waste (MSW) to a stable humus-like material known as compost. Applications of aerobic composting include (1) yard waste, (2) separated MSW, (3) commingled MSW, and (4) co-compo sting with wastewater sludge. Process descriptions and design guidelines for aerobic composting are presented in this section.

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3.2.1 Process description All aerobic composting processes are similar in that they all incorporate three basic steps: (1) preprocessing of the MSW, (2) aerobic decomposition of the organic fraction of the MSW, and (3) product preparation and marketing. Windrow, aerated static pile, and in-vessel are the three principal methods used for the composting of the organic fraction of MSW. While these processes differ primarily in the method used to aerate the organic fraction of solid waste, the biological principles remain the same, and, when designed and operated properly, all produce a similar-quality compost in approximately the same time period. 3.2.2 Process microbiology During the aerobic composting process a succession of facultative and obligate aerobic microorganisms is active. In the beginning phases of the composting process, mesophilic bacteria are the most prevalent. After the temperatures in the compost rise, thermophilic bacteria predominate, leading to thermophilic fungi, which appear after 5 to 10 days. In the final stages, or curing period as it is sometimes known, actinomycetes and molds appear. Because significant concentrations of these microorganisms may not be present in some types of biodegradable waste (e.g., newspaper), it may be necessary to add them to the composting material as an additive or inoculum. The microbiology of all aerobic composting processes is similar. Critical parameters in the control of aerobic composting processes include moisture content, C/N ratio, and temperature. For most biodegradable organic wastes, once the moisture content is brought to a suitable level (50 to 60 percent) and the mass aerated, microbial metabolism speeds up. The aerobic microorganisms, which utilize oxygen, feed upon the organic matter and develop cell tissue from nitrogen, phosphorus, some of the carbon, and other required nutrients. Much of the carbon serves as a source of energy for the organisms and is burned up and respired as carbon dioxide. Because organic carbon can serve both as a source of energy and cell carbon, more carbon is required than, for example, nitrogen. 3.2.3 Design and operational considerations Particle size Most materials that comprise the organic fraction of MSW tend to be irregular in shape. This irregularity can be reduced substantially by shredding the organic materials before they are composted. Particle size influences the bulk density, internal friction and flow characteristics, and drag forces of the materials. Most important of all, a reduced particle size increases the biochemical reaction rate during aerobic composting process. The most desirable particle size for compo sting is less than 2 inches (5 cm), but larger particles can be composted. The particle size of the material being composted is governed to some extent by the finished-product requirements and by economic considerations. Carbon-to-nitrogen ratio The most critical environmental factor for composting is the carbon-to-nitrogen ratio (C/N ratio). The optimum range for most organic wastes is from 20 to 25 to 1. As shown in Table 3.1, sludges have low C/N ratios, whereas yard wastes, such as leaves and newspaper, have relatively high C/N ratios. It should be noted that the C/N ratios given in Table 3.1 are based on the total dry weights of carbon and nitrogen, not on the dry weight of the biodegradable fraction of the organic material. In general, all of the organic nitrogen present in most organic compounds will become available, whereas not all of the organic carbon will be biodegradable. Depending on the

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particular waste material, the C/N ratio computed on the basis of total weights of carbon and nitrogen could be quite misleading, especially in those cases where all of the available nitrogen is biodegradable, but only a portion of organic carbon is biodegradable, (e.g., lignin in waste paper). For example, assuming that all of the nitrogen is available, the C/N ratio for the organic fraction of MSW can vary from about 34 to 60 depending on whether it is assumed that the available carbon is partially or totally biodegradable. Blending of a waste high in carbon and low in nitrogen (e.g., newsprint) with a waste that is high in nitrogen (e.g., yard wastes) is used to achieve optimum C/N ratios for composting. Blending and seeding Two design factors that may affect the blending of organic fraction of municipal solids waste for composting are C/N ratio and moisture content. Laboratory analyses are usually required to determine how the various organic materials should be blended for aerobic composting. If the organic fraction of MSW contains significant amounts of paper or other substrates rich in carbon, other organic materials such as yard wastes, manure, or sludge from wastewater treatment plants can be blended to provide a near optimum C/N ratio. Similarly, materials too wet and too dry for good composting can be blended in proper proportion to achieve an optimum moisture content. Seeding involves the addition of a volume of microbial culture sufficiently large to effect the decomposition of the receiving material at a faster rate. Table 3.1 Nitrogen content and nominal C/N ratios of selected compostable material (dry weight)

Source: Tchobanoglous et al., 1993.

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Moisture content The optimum moisture content for aerobic compo sting is in the range of 50 to 60 percent. Moisture can be adjusted by blending of components or by addition of water. When the moisture content of compost falls below 40 percent, the rate of composting will be slowed. Mixing/turning Initial mixing of organic wastes is essential to increase or decrease the moisture content to an optimum level. Mixing can be used to achieve a more uniform distribution of nutrients and microorganisms. Thrning of the organic material during the composting process is a very important operational factor in maintaining aerobic activity. Because turning can be specified by moisture content, waste characteristics, or air requirements, it is impossible to specify a minimum frequency of turning or number of turns in a general terms. For an organic waste having a maximum moisture of 55 to 60 percent and a compo sting period of 15 days, the first turn has been suggested at the third day. Thereafter, it should be turned every other day for a total of four to five turns. Temperature Aerobic composting systems can be operated in either the mesophilic, 30 to 38°C (85 to 100°F), or the thermophilic, 55 to 60°C (131 to 140°F), temperature regions. The temperature rise observed in actively composting wastes is caused by the exothermic reactions associated with respiratory metabolism. In aerated static pile and in-vessel composting systems, the temperature can be regulated by monitoring the temperature and controlling the airflow. In windrow composting, temperature can only be controlled indirectly, by varying the frequency of turning based on temperature measurements. In general, pile temperature will drop 5 to 10°C after turning, but will return to its previous level within several hours. Windrow temperatures decrease after 10 to 15 days as the readily biodegradable organic material is oxidized. Control of pathogens The destruction of pathogenic organisms is an important design element in the compost process, as it will affect the temperature profile and aeration process. Data on the thermal death points for a number of pathogenic organisms are summarized in Table 3.2. As shown in Table 3.2, the die-off of pathogens is a function of time and temperature. For example, the Salmonella species of bacteria can be destroyed in 15 to 20 minutes when exposed to a temperature of 60°C, or in one hour at 55°C. From Table 14-8, it is apparent that most pathogens will be destroyed rapidly when all parts of the compost pile are subjected to a temperature of about 55°C. Only a few can survive at temperatures up to 67°C for a short period of time. Elimination of all pathogenic microorganisms can be accomplished by allowing the composting waste to reach a temperature of 70°C for 1 to 2 hours. The U.S. Environmental Protection Agency has required specific time-temperature standards for pathogen control in composting systems (see Table 3.3). These conditions are easily met in properly operating composting systems. Air requirements In processes with forced aeration, such as the aerated static pile and the in-vessel systems, the total air requirement and air flow rate are essential design parameters. pH control

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Control of pH is another important parameter in evaluating the microbial environment and waste stabilization. The pH value, like the temperature, of the compost varies with time during the composting process. The initial pH of the organic fraction of MSW is typically between 5 and 7. The pH of compo sting material will vary according to the pH-time profile shown in Fig. 14-3a. In the first few days of composting, the pH drops to 5 or less. At this stage, the organic mass is at ambient temperature, the multiplication of indigenous mesophilic organisms begin, and the temperature rises rapidly. Among the products of this initial stage are simple organic acids, which cause the drop in pH. After about three days, the temperature reaches a thermophilic stage, and the pH begins to rise to approximately 8 or 8.5 for the remainder of the aerobic process. The pH value falls slightly during the cooling stage and reaches to a value in the range of 7 to 8 in the mature compost. If the degree of aeration is not adequate, anaerobic conditions will occur, the pH will drop to about 4.5, and the composting process will be retarded. Table 3.2 Temperature and time of exposure required for destruction of some common pathogens and

parasites


Table 3.3 EPA requirements for pathogen control in compost processes

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Degree of decomposition A suitable methodology for the measurement of the degree of decomposition is not available. However, several methodologies have been proposed. The proposed methods are (a) final drop in temperature, (b) degree of self-heating capacity, (c) amount of decomposable and resistant organic matter in the composted material, (d) rise in the redox potential, (e) oxygen uptake, (f) growth of the fungus Chaetomium gracilis, and (g) the starch-iodine test. The laboratory analysis of chemical oxygen demand (COD) and the lignin test provide a quick check for determining the degree of decomposition. A low COD value and a high lignin content (greater than 30 percent) is indicative of a stable compost. Control of odor The majority of the odor problems in aerobic composting processes are associated with the development of anaerobic conditions within the compost pile. In many large-scale aerobic composting systems, it is common to find pieces of magazines or books, plastics (especially plastic films), or similar materials in the organic material being composted. These materials normally cannot be decomposed in a relatively short time in a compost pile. Furthermore, because sufficient oxygen is often not available in the center of such materials, anaerobic conditions can develop. Under anaerobic conditions, organic acids will be produced, many of which are extremely odorous. To minimize the potential odor problems, it is important to reduce the particle size, remove plastics and other nonbiodegradable materials from the organic material to be composted, or use source-separated or uncontaminated feedstocks. Land requirements Land area requirements are another important element which must be considered in the aerobic composting processes. For example, in windrow compo sting for a plant with a capacity of 50 ton/d, about 2.5 acres of land would be required. Of this total, 1.5 acres would be devoted to buildings, plant equipment, and roads. For each additional 50 tons, it is estimated that 1.0 acre would be required for the composting operation and that 0.25 acre would be required for buildings and roads. The land requirement for highly mechanized systems varies with the process. An estimate of 1.5 to 2.0 acres for a plant with a capacity of 50 ton/d is not unreasonable; for larger plants, the unit area requirements would be less. Processing compost for market The economics of compost systems are greatly enhanced if the compost can be sold. To be marketable, compost must be of a consistent size; free from contaminants such as glass, plastic, and metals; and free of objectionable odors. The type of processing used to prepare the compost for marketing will depend on the specifications for the compost. Shredding and screening are commonly used to produce a more uniform product. In some cases, additives may be added to enhance the value of the final product. 3.2.3 Selection of aerobic composting processes Because the performance of properly operating windrow, aerated static pile, and in-vessel composting processes is essentially the same, the selection among alternative processes is based on capital and operating costs, land availability, operational complexity, and potential for nuisance problems. An additional factor to consider is that the windrow and aerated static pile processes utilize standard components and have been designed and constructed successfully "in-house" by many

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municipalities and private firms. The in-vessel processes employ proprietary designs and custom-made equipment and are usually procured on a turnkey basis from a single vendor. Many in-vessel compost systems are procured on full-service contracts and are owned and operated by the vendor or a third party. The economics of these arrangements should be compared carefully to ownership and operation of a simpler windrow or aerated static pile system by the municipality itself. A successful composting operation is highly dependent on proper operation and maintenance as well as design.

3.3 LOW-SOLIDS ANAEROBIC DIGESTION Low-solids anaerobic digestion is a biological process in which organic wastes are fermented at solids concentrations equal to or less than 4 to 8 percent. The low-solids anaerobic fermentation process is used in many parts of the world to generate methane gas from human, animal, and agricultural wastes, and from the organic fraction of MSW. One of the disadvantages of the low-solids anaerobic digestion process as applied to solid wastes is that considerable water must be added to wastes to bring the solids content to the required range of 4 to 8 percent. The addition of water results in a very dilute digested sludge, which must be dewatered prior to disposal. The disposal of the liquid stream resulting from the dewatering step is an important consideration in the selection of the low-solids digestion process. 3.3.1 Process description There are three basic steps involved whenever the low-solids anaerobic digestion process is used to produce methane from the organic fraction of MSW. The first step involves the preparation of the organic fraction of the MSW. Typically, for commingled solid waste the first step involves receiving; sorting and separation; and size reduction. Size reduction is also required for source-separated materials. The second step involves the addition of moisture and nutrients, blending, pH adjustment to about 6.8, and heating of the slurry to between 55 and 60°C, and the anaerobic digestion is carried out in a continuous-flow reactor whose contents are mixed completely. In some operations, a series of batch reactors have been used instead of one or more continuous-flow completemix reactors.

In most operations, the required moisture content and nutrients are added to the wastes to be processed, in the form of wastewater sludge or cow manure. Depending on the chemical characteristics of the sludge or manure, additional nutrients may also have to be added. Because foaming and the formation of surface crusts have caused problems in the digestion of solid wastes, adequate mixing is of fundamental importance in the design and operation of such systems. The third step in the process involves the capture, storage, and, if necessary, separation of the gas components. The dewatering and disposal of the digested sludge is an additional task that must be accomplished. In general, the processing of the digested sludge produced from low-solids anaerobic digestion is so expensive that the process has seldom been used.

3.3.2 Process microbiology Carried out in the absence of oxygen, the anaerobic stabilization process or conversion of the organic materials in MSW occurs in three steps. As discussed previously, the first step in the process involves the enzyme-mediated transformation (hydrolysis) of higher-molecularmass compounds into compounds suitable for use as a source of energy and cell tissue. The second

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step involves the bacterial conversion of the compounds resulting from the first step into identifiable lower-molecular-mass intermediate compounds. The third step involves the bacterial conversion of the intermediate compounds into simpler end products, principally methane and carbon dioxide. 3.3.3 Process design considerations Although the process of anaerobic digestion of the organic fraction of MSW is not developed fully, some of the important design considerations are summarized in Table 3.4. In operations where solid wastes have been mixed with wastewater sludge, it has been found that the gas collected from the digesters contains between 50 and 60 percent methane. It has also been found that about 10 ft3 of gas is produced per pound of biodegradable volatile solids destroyed. Because of the variability of the results reported in the literature, it is recommended that pilot plant studies be conducted if the digestion process is to be used for the conversion of MSW or other organic wastes. Process selection Process selection between anaerobic processes is typically between the low-solids process and the high-solids process, described in the following section. Selection of equipment and facilities for the low-solids anaerobic digestion process usually involves the type of mixing equipment (internal mixers, internal gas mixing, and external pump mixing), the general shape of the digester (e.g., circular or egg-shaped), the control systems, and the ancillary facilities needed for mixing the incoming wastes and dewatering the digested sludge. Table 3.4 Important design considerations for the low-solids anaerobic digestion of the organic

fraction of MSW


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3.4 HIGH-SOLIDS ANAEROBIC DIGESTION

High-solids anaerobic digestion is a biological process in which the fermentation occur at a total solids content of about 22 percent or higher. The high-solids anaerobic digestion is a relatively new technology and its application for energy recovery from the organic fraction of MSW has not been developed fully. Two important advantages of the high-solids anaerobic digestion process are lower water requirements and higher gas production per unit volume of the reactor size. The major disadvantage of this process is that at present (1992), limited full-scale operating experience is available. 3.4.1 Process description The three steps described for the low-solids anaerobic digestion are also applied in high-solids anaerobic digestion process. The principal difference is at the end of the process, where less effort is required to dewater and dispose of the digested sludge.

3.4.2 Process microbiology The process microbiology for the high-solids anaerobic digestion is as described previously for the low-solids anaerobic digestion process. However, because of the high solids concentration, the effects of many environmental parameters on microbial population are more severe. For example, ammonia toxicity can affect the methanogenic bacteria, which will have an adverse effect on system stability and methane production. In most cases, the ammonia toxicity can be prevented by a proper adjustment of the C/N ratio of the input feedstock.

3.4.3 Process design considerations Although the high-solids anaerobic digestion process is not developed fully, some of the important design considerations are summarized in Table 3.5. In general, the high-solids anaerobic digestion process is capable of stabilizing more organic waste and producing more gas per unit of volume of reactor than is the low-solids process considered previously. Table 3.5 Important design considerations for the high-solids anaerobic digestion of the organic fraction

of MSW

Item Comment Size of material Wastes to be digested should be shredded to a size that will not

interfere with the efficient functioning of feeding and discharging mechanisms

Mixing equipment The mixing equipment will depend on the type of reactor to be used Percentage of sold wastes mixed with sludge

Depends on the characteristics of the sludge

Mass retention time Use 20 to 30 for design, or base design on results of pilot plant studies Loading rate based on biodegradable volatile solids

6 to 7 kg/m3.d. Not well defined at present time. Significantly higher rates have been reported

Solid concentration Between 20 and 35% (22 to 28% typical) Temperature Between 30 and 38oC for mesophilic and between 55 and 60oC for

thermophilic reactor Gas production 0.625 to 1.0 m3/kg of biodegradable volatile solids destroyed (CH4 = 50

percent; CO2 = 50 percent) Source: Tchobanoglous et al., 1993.

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3.4.4 Process selection At the present time, as described in the following section, there are no full scale high-solids digestion processes in operation in the United States. However, there are several different high-solids processes under development in the United States and Europe, and at least three are operational in Europe. As these processes become more fully developed and tested, it is anticipated that a number of anaerobic processes will be available commercially. 3.5 DEVELOPMENT OF ANAEROBIC DIGESTION PROCESSES AND

TECHNOLOGIES FOR TREATMENT OF THE ORGANIC FRACTION OF MSW The purpose of this section is to introduce the reader to the emerging technologies involving the use of the anaerobic digestion process for the production of methane and a humus product from the organic fraction of MSW and other organic materials. 3.5.1 Anaerobic digestion technologies In recent years, there has been a great interest in applying the anaerobic digestion process for the processing of the organic fraction of MSW because of the opportunity to recover methane and the fact that the digested material is similar to compost produced aerobically. Most of the work with the anaerobic digestion process is going on in Europe. The combined high-solids anaerobic digestion/aerobic composting process under development in the United States is considered below. 3.5.2 Combined high-solids anaerobic digestion/aerobic composting The high-solids anaerobic digestion/aerobic composting process, developed by Professor Bill Jewell at Cornell University, combines the high-solids anaerobic digestion and aerobic composting processes. The major advantage of this process is the complete stabilization of the organic waste with a net energy recovery and without the need for major dewatering equipment. Other advantages include pathogen control and volume reduction. Process description The high-solids anaerobic digestion/aerobic composting process is a two-stage process. The first stage of the two-stage process involves the high-solids (25 to 30 percent) anaerobic digestion of the organic fraction of MSW to produce a gas composed of methane and carbon dioxide. The anaerobic reactor operates under thermophilic conditions 129 to 133°F (54 to 56°C) with a nominal hydraulic retention time of 30 days. The second stage involves the aerobic composting of the anaerobically digested solids to increase the solids content from 25 to 65 percent or more, depending on the final use. The output from the second stage is a fine humus-like material with a thermal content of about 6000 to 6400 Btu/lb and a specific weight of about 35 lb/ft3. Because the final humus that is produced is combustible, it appears that it can be fired directly in a boiler when mixed with other fuels or pelletized for use as a fuel source. Alternatively, the humus-like material can be used as a soil amendment.

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Process applications The combined high-solids anaerobic digestion/aerobic composting process is in the early stages of development. Design considerations for the first stage of this process are the same as for the high-solids anaerobic digestion process. The two major design parameters for the second stage are the aeration and thermal energy requirement for the destruction of pathogens. This process can be used to process the combined organic fraction of MSW and wastewater treatment plant sludge. Because the highsolids anaerobic digestion/aerobic compostion process can be used to process both the organic fraction of MSW along with wastewater treatment plant sludge, the use of costly sludge dewatering facilities and the need to treat the liquid resulting from the dewatering of the sludge can be eliminated. Biogas produced from this process can be used for methanol production. The heat required for the anaerobic and aerobic reactors will be recovered from the thermal energy of the fluidized bed combustion reactor.

3.6 OTHER BIOLOGICAL TRANSFORMATION PROCESSES The principal biological processes used for the transformation of the organic fraction of MSW are summarized in Table 3.6. Apart from the aerobic composting and the anaerobic digestion processes, enzymatic hydrolysis and fermentation following acid or enzymatic hydrolysis are the biological processes that have received most attention. With the national focus on solid waste management and increasing landfill disposal costs, it is anticipated that a number of new processes will become available in the future. Table 3.6 Biological processes for the recovery of conversion products from the organic fraction of MSW

Process Conversion product Preprocessing Aerobic conversion Compost (soil conditioner) Separation of organic fraction,

particle size reduction Anaerobic digestion (in landfill) Methane and carbon dioxide None, other than placement in

containment cells Anaerobic digestion (low-solids, 4 to 8 percent solids)

Methane and carbon dioxide, digested solids

Separation of organic fraction, particle size reduction

Anaerobic digestion (high-solids, 22 to 35 percent solids)

Methane and carbon dioxide, digested solids

Separation of organic fraction, particle size reduction

Enzymatic hydrolysis Glucose from cellulose Separation of cellulose-containing materials

Fermentation (following acid or enzymatic hydrolysis)

Ethanol, single-cell protein Separation of organic fraction, particle size reduction, acid or enzymatic hydrolysis to produce glucose


3.7 CHEMICAL TRANSFORMATION PROCESSES Chemical transformation processes include a number of hydrolysis processes, which are used to recover compounds such as glucose and furfural, and a variety of other chemical conversion processes used to recover compounds such as synthetic oil, gas, and cellulose acetate. Methanol, an alternative liquid fuel, can also be produced. These chemical processes are not used routinely for the transformation of the organic fraction of MSW, because these compounds can also be manufactured from other cellulose-containing wastes, such as wheat straw, sugar cane bagasse, and corncobs. The economic viability of these processes is closely linked to the cost of alternative feedstocks. For example, agricultural wastes are currently cheaper to procure than either source-separated or machineprocessed MSW.

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Acid hydrolysis The cellulose molecule is comprised of about 3000 glucose units, is soluble in water and many organic solvents, and is relatively immune to attack by most microorganisms. If the cellulose molecule is hydrolyzed, the glucose can be recovered. Acid hydrolysis, used to recover glucose from cellulose, involves treating a finely divided suspension of cellulose-containing waste (e.g., newsprint) with a weak acid. The suspension is then heated to between 180 and 230°C and slight pressure is applied. Under these conditions, the cellulose in the waste is converted into glucose and other sugars. The amount of glucose recovered depends on the characteristics of the waste. It is estimated that upwards of 80 percent of the weight of kraft paper may be recovered as sugar. acid (C6H10O5)n + H2O nC6H1206 cellulose glucose Lignin is not affected by the process. The sugar and glucose extracted from the cellulose can be converted by other chemical and biological processes into alcohols and other industrial chemicals. Methanol production from methane The methane produced by the anaerobic digestion of the organic portion of MSW can be converted to methanol, a liquid fuel. The conversion process involves the following two reactions, which are carried out in series. catalyst CH4 + H2O CO + 3H2 catalyst CO + 2H2 CH3OH In the first reaction, which is endothermic, biogas, containing methane, is reacted with steam in a catalyst filled reactor to form carbon monoxide and hydrogen gas. In the second reaction which is exothermic, the products of the first reaction are converted catalytically to form methanol. The principal advantage of producing methanol from biogas that contains methane is that the resulting fuel is both storable and transportable. Methanol is currently manufactured from natural gas at a lower cost than it could be from biogas produced from the anaerobic digestion of MSW. Because the cost of fossil fuels is highly sensitive to political trends, the relative economics of methanol production from biogas could change in the future.

3.8 ENERGY PRODUCTION FROM BIOLOGICAL CONVERSION PRODUCTS Once conversion products have been derived from solid wastes by either anaerobic digestion (methane), or chemical transformation (methanol), the next step involves their uses and/or storage. If energy is to be produced from these products, an additional conversion step is required. Biogas can be used directly with internal combustion (IC) engines and gas turbines to generate electricity. Where IC engines are used, they should be started with propane or natural gas and operated until the engine is at its operating tempeature, at which point they can be switched to biogas operation. Before an engine is shut down, it again should be operated with propane or natural gas for about 20 minutes or so. By starting and stopping the engine with propane or natural gas, corrosion problems associated with hydrogen sulfide can be avoided. Biogas from landfills is being used to fuel IC engines in a wide range of power outputs from 50 kW to 5 MW.

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With gas turbines, the biogas is compressed under high pressure so that it can be used more effectively in the turbine. Such systems are in widespread use fueled with biogas recovered from landfills. Gas turbine systems are generally used in the 1 to 5 MW power range. In large installations, the most common flow diagram for the production of electric energy involves the use of a steam turbine-generator combination. Steam can be produced in a boiler fired with either biogas (methane) or MSW derived liquid fuel (methanol). A number of steam turbine-generator installations are in operation thoughout the United States, fueled with biogas recovered from landfills. The largest such installation in the United States generates 50 MW of electricity at the Puente Hills landfill near Whittier, California.

Environmental Biotechnology lecture note-220110

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