15.1 Introduction 571
15.2 Sanitary Landfill 572
15.3 Leachate 573
15.4 Composition and Characteristics of Leachate 573
15.4.1 Leachate of Different Age 573
15.4.2 Leachate in Different Countries 573
15.5 Leachate Treatment 574
15.6 Bioremediation Methods 574
15.6.1 In Situ and Ex Situ Methods 574
15.6.2 Advantages and Disadvantages of Bioremediation 575
15.6.3 Physiology of Biodegradative Microbes 575
15.6.4 Metabolic Processes 576
15.6.5 Factors Affecting Bioremediation 576
15.7 Bioremediation of Landfill Leachate 577
15.8 Case Studies 580
15.8.1 Case 1: Anaerobic/Aerobic Treatment of Municipal Landfill Leachate in Sequential Two-Stage Up-Flow Anaerobic Sludge Blanket Reactor
(UASB)/Aerobic Completely Stirred Tank Reactor (CSTR) Systems 580
15.8.2 Case 2: Comparison of Two Biological Treatment Processes Using Attached-Growth Biomass for Sanitary Landfill Leachate Treatment 581
15.8.3 Case 3: Leachate Treatment Using an Aerobic Biofilm Reactor 583
More than 90% of municipal solid waste is directly disposed of on land, the vast majority of it in an unsatisfactory manner. Open and burning dumps are common in many developing countries; these contribute to water and air pollution and provide food and breeding grounds for birds, rats, insects, and other carriers of disease. The presence of these dumps often reduces the property value of nearby land and residences.
Sanitary landfilling is an acceptable and recommended method for ultimately disposing of solid wastes. This method has sometimes been confused with waste disposal on open and burning dump sites, but this is a misconception. The sanitary landfill is an engineered landfill that requires sound and detailed planning and specification, careful construction, and efficient operation. In essence, modern landfilling involves spreading the wastes in thin layers, compacting them to the smallest practical volume, and covering them with daily earth cover in a manner that minimizes adverse environmental pollution.
The sanitary landfill, the most acceptable alternative to the present poor practices of land disposal, involves the long-term planning and application of sound engineering principles and construction techniques. By definition, no burning of solid waste will ever occur at a sanitary landfill. A sanitary landfill is not only an acceptable and economic method of solid waste disposal, it is also an excellent way to make otherwise unsuitable or marginal land valuable.1
All landfills produce a liquid stream called leachate, which is a highly complex and polluted wastewater. Leachate pollution is a concern for many local authorities as it directly degrades river water quality. Many researchers continue to search for ways to treat leachate effectively using different biological processes. To secure long-term dewatering of landfills and reduce the increasing treatment costs, it is therefore necessary to control leachate quantity and quality. This is often difficult, as increasing water quality standards make the requirements on leachate treatment ever more stringent.
Treatment procedures must consider the highly varying flow and complex composition of the leachate; this often results in special operational problems. The following chapters give an overview of leachate generation and the development of leachate control and treatment applicable to many landfills.
A sanitary landfill is defined as a land disposal site that applies an engineered method of disposing of solid wastes on land in a manner that minimizes environmental hazards by spreading the solid wastes to the smallest practical volume, and applying and compacting cover material at the end of each day.2
Landfills are the physical facilities used for the ultimate disposal of residual solid wastes in the ground. In the past, the term sanitary landfill was used to denote a landfill in which the wastes were placed in the landfill and then covered at the end of daily operation. Today, sanitary landfill refers to an engineered facility for the disposal of municipal solid waste (MSW), designed and operated to minimize public health and environmental impact.
Solid wastes deposited in a landfill undergo slow degradation to produce residual solid, liquid, and gaseous products. Ferrous and other metals are oxidized and organic and inorganic wastes are utilized by microorganisms through aerobic and anaerobic processes. Organic acids, which are produced as a result of microbial degradation, increase chemical activity within the fill. Food wastes degrade quite readily, but other materials, such as plastics, rubber, glass, and some demolition wastes, are highly resistant to decomposition.
The degree of degradation of organic waste in landfills is very much dependent on the organic content of the waste. Wastes in Asian countries are reported to have a larger organic fraction, which leads to more problems in leachate generation. Waste data from Indonesia and China show that the organic fraction comprised 70.2% and 67.3%, respectively.3
Landfill methods are considered the most economical and environmentally acceptable way of disposing of solid wastes throughout the world. Even with the implementation of waste reduction, recycling, and transformation technologies, disposal of residual solid waste in landfill will still remain an important component of an integrated solid waste management strategy.4
In engineering terms, a sanitary landfill is also sometimes identified as a bioreactor due to the presence of anaerobic activities in the wastes. As such, landfilling sites need the incoming waste stream top be monitored, as well as placement and compaction of the waste, and installation of landfill environmental monitoring and control facilities. Gas vent and leachate collection pipes are important features of a modern landfill.
The harmful liquid that collects at the bottom of a landfill is known as leachate. The generation of leachate is a result of uncontrolled runoff, and percolation of precipitation and irrigation water into the landfill. Leachate can also include the moisture content initially contained in the waste, as well as infiltrating groundwater. Leachate contains a variety of chemical constituents derived from the solubilization of the materials deposited in the landfill and from the products of the chemical and biochemical reactions occurring within the landfill under the anaerobic conditions.
An estimation of leachate generation in a landfill can be carried out by calculating the infiltration through a landfill cover using a water budget model such as the Hydrologic Evaluation of Landfill Performance (HELP).5 The model uses conservation of mass to predict water movement, which enables the volumetric flux of water infiltrating into the waste to be calculated on a time-varying basis.
The generated leachate can cause significant environmental damage, becoming a major pollution hazard when it comes into contact with the surrounding soil, ground, or surface waters. One such problem is caused by infiltrating rainwater and the subsequent movement of liquid or leachate out of the fill into the surrounding soil. This leachate often contains a high concentration of organic matter and inorganic ions, including ammoniacal nitrogen and heavy metals. Therefore, in order to avoid environmental damage, landfill leachate must be collected and appropriately treated before being discharged into any water body.
Leachate tends to percolate downward through solid waste, continuing to extract dissolved or suspended materials. In most landfills, leachate seeps through the landfill from external sources, such as surface drainage, rainfall, groundwater, and water from underground springs, as well as from the liquid produced from the decomposition of the wastes, if any.3
Many factors influence the production and composition of leachate. One major factor is the climate of the landfill. For example, where the climate is prone to higher levels of precipitation, there will be more water entering the landfill and therefore more leachate generated. Another factor is the site topography of the landfill, which influences the runoff patterns and again the water balance within the site.
The composition of leachate is important in determining its potential effects on the quality of nearby surface water and groundwater. Contaminants carried in leachate are dependent on solid waste composition and on the simultaneously occurring physical, chemical, and biological activities within the landfill. The quantity of contaminants in leachate from a completed landfill where no more waste is being disposed of can be expected to decrease with time, but it will take several years to stabilize.
The decomposition of solid urban waste in landfills is essentially a result of microbiological processes and, therefore, the production of biogas and leachate are both directly related to the activity of microorganisms. It has been demonstrated that large variations in leachate quality exist for different landfills, but also at different locations within the same landfill.6
New landfills generate more organic pollutants than older landfills. The BOD : COD (biochemical oxygen demand : chemical oxygen demand) ratio in young leachate is typically in the range of 0.5 to 0.7, which indicates higher biodegradability than that of mature landfills, which produce leachate with a BOD : COD ratio of less than 0.4.
It is expected that leachate characteristics will vary by country. This is because the soil under a landfill site, the composition of disposed waste, the climate, sampling and landfill management vary among countries.7,8
Many landfills pollute water bodies by discharging untreated leachate. When leachate percolates through the ground, it entrains landfill components such as decaying organic matter, microorganisms, metals, and inorganic compounds into the underlying groundwater, causing serious contamination.
Landfill leachates are commonly classified as a high-strength wastewater containing dissolved and entrained landfill components.9 Freshly produced landfill leachates are characterized by low pH values, high BOD5 and COD values, as well as by the presence of several other toxic/hazardous compounds.10 Several treatment options have been utilized for leachate treatment, with varying degrees of efficiency. The main applicable methods are biological, chemical, membrane separation, and thermal treatment processes.11
Physico-chemical processes are generally considered to incur high operating costs and sometimes lower effectiveness. A biological process is normally preferred, such as a conventional activated sludge process, which has been proven to be effective for the removal of organic carbon and nutrient content. Nevertheless, the problem of poor sludge settleability has usually been encountered, as well as the need for longer aeration times, for settling tanks of larger volume, and for total biomass recycling.
Some landfills practice leachate recycling in the fill area, where leachate percolates through the waste cell and undergoes further degradation. The treatment process or processes selected will depend to a large extent on the contaminants to be removed.4
Biological processes have been increasingly used in the treatment of leachate in combination with physical and chemical processes. Selected microorganisms are introduced in the aerobic treatment to achieve a better process efficiency. However, because of the variation in leachate composition from site to site, the remedial process train will generally be tailored to the site and consist of several unit operations. The following section discusses applications of bioremediation processes to landfill leachates. It is important to remember that characterization of leachate plumes through groundwater modeling, analysis of leachate physical and chemical characteristics, and development of leachate recovery systems are all important in selecting a leachate treatment system.12
Bioremediation is defined as the use of microorganisms or microbial processes to degrade environmental contaminants. Bioremediation has numerous applications, including cleanup of groundwater, soils, lagoons, sludge, and process waste streams.
In general terms, bioremediation involves multiphase but heterogeneous environments, such as soils in which the contaminant is present in association with the soil particles, dissolved in soil liquids, and in the soil atmosphere. Because of these complexities, successful bioremediation depends on an interdisciplinary approach involving microbiology, biochemistry, and engineering.
For leachate treatment, the bioremediation method may be carried out either on or off site. Both methods have their advantages and disadvantages, depending on the site condition. Two factors favor the treatment of leachates on site: the expense of off-site transportation and the reluctance of communities nationwide to permit transportation routes or treatment facilities within their jurisdictions. The desirability of on-site leachate treatment should encourage the development of small-scale technology requiring low capital investment. Biological processes are well suited to on-site leachate treatment for the removal of organic compounds.9
Depending on the situation, a bioremediation method can be either ex situ or in situ. Ex situ treatments are treatments that involve the physical removal of the contaminated material in order to undergo the treatment process. In situ techniques involve treatment of the contaminated material in place. Examples of in situ and ex situ bioremediation are listed in the following:
1. Land farming. This is a solid-phase treatment system for contaminated soils; it may be carried out in situ or ex situ.
2. Bioreactors. Biodegradation is carried out in a container or reactor; it may be used to treat liquids or slurries.
3. Composting. This is an aerobic, thermophilic treatment process in which contaminated material is mixed with a bulking agent; it can be carried out using static piles or aerated piles.
4. Bioventing. This is a method of treating contaminated soils by drawing air or oxygen through the soil to stimulate microbial activity.
5. Biofilters. Microbial stripping columns are used to treat air or liquid emissions.
6. Bioaugmentation. Bacterial cultures are added to a contaminated medium; this is frequently used in both in situ and ex situ systems.
7. Biostimulation. Indigenous microbial populations in soils or groundwater are stimulated by providing the necessary nutrients.
8. Intrinsic bioremediation. This is the unassisted bioremediation of the contaminant; the only process carried out is regular monitoring.
15.6.2 Advantages and Disadvantages of Bioremediation
Successful bioremediation requires microbes and suitable environmental factors for degradation to occur. The most suitable microbes are bacteria or fungi that have the physiological and metabolic capabilities to degrade the pollutants.
Bioremediation offers several advantages over conventional methods of waste treatment such as landfilling or incineration. Bioremediation can be done on site, it is often less expensive, involves minimal site disruption, eliminates waste permanently, eliminates long-term liability, has greater public acceptance with regulatory encouragement, and can be coupled with other physical or chemical treatment methods.
Bioremediation also has its limitations. Some chemicals are not amenable to biodegradation, for instance, heavy metals, radionuclides, and some chlorinated compounds. In some cases, the microbial metabolism of the contaminants may produce toxic metabolites. Bioremediation is a scientifically intensive procedure that must be tailored to site-specific conditions, and usually requires treatability studies to be conducted on a small scale before the actual cleanup of a site.13 The treatability procedure is important, as it establishes the extent of degradation and evaluates the potential use of a selected microorganism for bioremediation. A precise estimate on vessel size or area involved, speed of reaction, and economics can therefore be determined at the laboratory stage.
Bioremediation is based on the activities of aerobic or anaerobic heterotrophic microorganisms. Microbial activity is affected by a number of physicochemical environmental parameters. Factors that directly affect bioremediation are energy sources (electron donors), electron acceptors, nutrients, pH, temperature, and inhibitory substrates or metabolites. One of the primary distinctions between surface soils, subsurface soils, and groundwater sediments is the organic material content. Surface soils, which typically receive regular inputs of organic material from plants, will undoubtedly have higher organic matter content.
High organic matter content is typically associated with high microbial numbers and a great diversity of microbial populations. Organic matter serves as a wardrobe of carbon and energy as well as a source of other macronutrients such as nitrogen, phosphorous, and sulfur. Subsurface soils and groundwater sediments have lower levels of organic matter and thus lower microbial numbers and population diversity than surface soils.14 Bacteria become more dominant in the microbial community with increasing depth in the soil profile, because the numbers of other organisms such as fungi or actinomycetes decrease. This is attributed to the ability of bacteria to use alternative electron acceptors to oxygen. Other factors that control microbial populations are moisture content, dissolved oxygen, nutrient, and temperature.13
The primary metabolism of an organic compound uses a substrate as a source of carbon and energy. For the microorganism, this substrate serves as an electron donor, which results in the growth of the microbial cell. The application of co-metabolism for bioremediation of a xenobiotic is necessary because the compound cannot serve as a source of carbon and energy due to the nature of the molecular structure, which does not induce the required catabolic enzymes. Co-metabolism has been defined as the metabolism of a compound that does not serve as a source of carbon and energy or as an essential nutrient, and can be achieved only in the presence of a primary (enzyme-inducing) substrate.
Two conditions favor metabolic activities: aerobic and anaerobic environments. Aerobic processes are characterized by metabolic activities involving oxygen as a reactant. Dioxygenases and monooxygenases are two primary enzymes used by aerobic organisms during the transformation and mineralization of xenobiotics. Anaerobic microbes take advantage of a range of electron acceptors, including, depending on their availability and the prevailing redox conditions, nitrate, iron, manganese, sulfate, and carbon dioxide.
15.6.5 Factors Affecting Bioremediation
The primary factor that affects the activity of bacteria is the ability and availability of reduced organic material to serve as an energy source. Whether a contaminant will serve as an effective energy source for an aerobic heterotrophic organism is a function of the average oxidation state of the carbon in the material. Each degradation process depends on microbial (biomass concentration, population diversity, and enzyme activities), substrate (physico-chemical characteristics, molecular structure, and concentration), and a range of environmental (pH, temperature, moisture content, availability of electron acceptors, and carbon and energy sources) factors. These parameters affect the acclimation period of the microbes to the substrate. Molecular structure and contaminant concentration have been shown to strongly affect the feasibility of bioremediation and the type of microbial transformation occurring, as well as whether the compound will serve as a primary, secondary, or co-metabolic substrate.
The rate at which microbial cells can convert contaminants during bioremediation depends on the rate of contaminant uptake and metabolism and the rate of transfer to the cell (mass transfer). Increased microbial conversion capacities do not lead to higher biotransformation rates when mass transfer is a limiting factor.15 This appears to be the case in most contaminated soils and sediments. For example, contaminating explosives in soil did not undergo biodegradation even after 50 years. Treatments involving rigorous mixing of the soil and breaking up of the larger soil particles stimulated biodegradation drastically.16 The bioavailability of a contaminant is controlled by a number of physico-chemical processes such as sorption and desorption, diffusion, and dissolution. Slow mass transfer causes a reduced bioavailability of the contaminants in the soil to the degrading microbes. Contaminants become unavailable when the rate of mass transfer is zero. The decrease of bioavailability over the course of time is often referred to as aging or weathering. It may result from the following:
1. Chemical oxidation reactions incorporating contaminants into natural organic matter
2. Slow diffusion into very small pores and absorption into organic matter
3. The formation of semirigid films around nonaqueous-phase liquids (NAPL) with a high resistance to NAPL-water mass transfer
These bioavailability problems may be overcome by the use of food-grade surfactants,17 which increase the availability of contaminants for microbial degradation.
Bioactivity refers to the operating state of microbiological processes. Improving bioactivity implies that system conditions are adjusted to optimize biodegradation.18 For example, if the use of biore-mediation requires meeting a certain minimum rate, adjusting the conditions to improve biodegradation becomes important and a bioremediation configuration that makes this control possible has an advantage over one that does not.
In nature, the ability of organisms to convert contaminants to both simpler and more complex molecules is very diverse. In light of our current limited ability to measure and control biochemical pathways in complex environments, favorable or unfavorable biochemical conversions are evaluated in terms of whether individual or groups of parent compounds are removed, whether increased toxicity is a result of the bioremediation process, and sometimes whether the elements in the parent compound are converted to measurable metabolites. These biochemical activities can be controlled in an in situ operation when one can control and optimize the conditions to achieve a desirable result.
Bioremediation is the treatment of choice for mineralizing most organic compounds in landfill leachate.19 Mineralization is carried out by microorganisms, which can degrade organic compounds to carbon dioxide under aerobic conditions and to a mixture of carbon dioxide and methane under anaerobic conditions. Microorganisms are also capable of changing the oxidation state of metals and inorganic compounds and can concentrate heavy metals and hydrophobic compounds through ingestion or adsorption. Microorganisms are ubiquitous, self-replicating, adaptable to a variety of leachate compositions, and active at moderate reaction conditions. In addition, biodegradation benefits from a long process history in the treatment of domestic sewage.
Leachate that comes from mixed landfills, that is, those with municipal waste combined with industrial wastes, may contain a host of xenobiotics (synthetic or unnatural) compounds. A number of these xenobiotics are normally classified as hazardous waste. A vast majority of organic hazardous wastes can be degraded if the proper microbial communities are established, maintained, and con-trolled.20 Degradation is not necessarily growth-associated,21 as organic compounds may be transformed to microbial storage polysaccharides under nitrogen-limiting conditions rather than being mineralized to carbon dioxide. Research regarding the mechanisms controlling xenobiotic degradation is important in understanding the capabilities and limitations of biological leachate treatment.22
An important element in xenobiotic biodegradation is the broad specificity of some microbial enzymes, which permits an enzyme-catalyzed reaction to occur without providing energy or carbon for cell replication. This phenomenon is divided into two categories: fortuitous metabolism, in which a growth co-substrate is not obligate, and co-metabolism, in which the growth co-substrate is obligate.23 One of the most thoroughly characterized examples of broad enzyme specificity is the ability of the methane mono-oxygenase enzyme (MMO) to oxygenate hydrocarbons other than methane, its natural substrate. The oxygenated hydrocarbons then accumulate stoichiometrically in the reactor.24 MMO-catalyzed reactions are co-metabolic, because energy from a co-substrate is required to supply reducing power for the reaction.
Fortuitous or co-metabolic biodegradation may account for a significant portion of the removal of xenobiotics in the environment.24 Numerous examples of co-metabolic activity have been described for pure substrates,22 but co-metabolism has been very difficult to demonstrate in mixed-substrate, mixed-culture systems, because products of the co-metabolic reactions of one species may be degraded by another.24 To encourage co-metabolism, easily degradable co-substrates should be included in the leachate prior to biological treatment. Fatty acids, which often occur in landfill leachates, may fulfill this requirement.
In the case of industrial landfill leachate, it is unlikely that the microbial enzymatic machinery would be sufficient to degrade all the compounds present,25 especially if a single microbial species is used. Furthermore, the adaptability of a single microbial species is limited and the mutation rate is too slow to make single-species adaptation practical. In order to increase the diversity of degradative enzymes it is common to use a mixed microbial population, also known as a microbial consortium or mixed culture. Mixed cultures have two advantages over pure cultures in the degradation of complex substrates. First, the product of an incomplete mineralization by one microbe, such as from a co-metabolic transformation, may serve as a substrate for another microbe. Second, the transfer of genetic information between species may enhance the degrad-ability of the culture.26 It has been demonstrated that DDT (dichloro diphenyl trichloroethane) can be co-metabolized to pentachlorophenol-induced periplasmic protein (PCPA) by one species and that PCPA can be mineralized by another species. A combined culture of the two species results in the complete mineralization of DDT.27 Stable mixed cultures degrading xenobiotics have been isolated in which the microbial consortia can degrade a substrate better than the individual species.22
Many strains of microorganism have been isolated that can degrade xenobiotics or families of xenobiotics.28 For example, a white rot fungus studied for its lignin-degrading potential has been shown in laboratory studies to mineralize a number of recalcitrant organics, such as a tetrachlorodibenzo-paradioxin (TCDD) and DDT.29 Degradation is carried out by extracellular enzymes whose production is stimulated by nitrogen limitation. Because of the requirements of nitrogen limitation and an acidic environment, the fungus is incompatible with many activated-sludge-derived organisms. Whether such organisms will be useful for degrading mixtures of compounds or will be active in a full-scale process has yet to be demonstrated.
Gross genetic changes brought about by the interspecies transfer of genetic material may be important in the microbial degradation of xenobiotics. Although there are several mechanisms for such transfers, the most important is thought to be conjugation. In this process, loops of extrachromosomal DNA mediate their own replication from host to recipient microorganisms. Conjugative plasmids, as these DNA loops are known, carry coding for a variety of proteins, which, although not required for reproduction, may confer a selective environmental advantage such as heavy metal resistance or extended substrate range.30 In some cases, nonconjugative plasmids can link to conjugative plasmids and "piggy-back" from organism to organism.23 Once a plasmid is transferred, DNA sequences called transposons may play a role in the integration of portions of the plasmid DNA into the genome of the new host. The rapid spread of antibiotic resistance among various classes of microorganisms is an example of the transfer of plasmid-born information.
The key issues in developing an effective biological landfill leachate treatment process are the following:
1. Process configuration
2. Microbial culture selection and development
3. Substrate modification
Due to the complex and varying nature of landfill leachate, these factors must be evaluated for each site. Chemical species thought to be biologically recalcitrant may be biodegradable given the proper acclimation. The principal mechanisms of acclimation are macromolecule modification, population selection, and genetic transfer. Modification of cellular components, for example, enzyme induction or increased membrane permeability, occurs when a substrate interacts with biological molecules of the cell. The time frame for such interactions ranges from minutes to hours.30 Population selection, or shifts in the representation of preexisting species, occurs because some species or mutants within a species may be better adapted to a new environment. The time frame depends on growth rates and may range from hours to days for aerobic cultures and from days to weeks for anaerobic cultures.31 Favorable genetic adaptation, alteration of the microbial DNA, may occur over periods ranging from months or years.32
Carbon limiting is also used to encourage enzyme induction, place the population under selective pressure for degradation of recalcitrant substrates, and favor the simultaneous rather than sequential metabolism of a mixed carbon source.33 Carbon-limiting conditions can be achieved either through continuous culture (chemostat) or through a fed batch reaction.
To facilitate biodegradation, the leachate may require modification through pH adjustment, removal or addition of oxygen, amendment with nutrients, or dilution or removal of toxic species. Microbial nutrition is complex and is better understood for aerobes than for anaerobes.34 Biological processes typically favor a pH near 7. Pretreatment processes to remove inhibitory components include coagulation and precipitation, carbon adsorption, and possibly ozonation.
A variety of biological processes options may be used to treat leachate.35 The basic decision is whether to treat a particular leachate aerobically or anaerobically. Both aerobic and anaerobic processes can degrade a wide range of xenobiotics.36 Aerobic processes are generally superior in mineralizing aromatic compounds; anaerobic processes are superior for short-chain aliphatic groups.27 Aerobic processes have the advantage of speed and ease of control and acclimation. However, aerobic processes accumulate large quantities of microbial sludge that may contain adsorbed organics and heavy metals, and may strip volatile compounds. Anaerobic processes produce less sludge and can provide energy through methane production. They also reduce sulfate to sulfide, which is a powerful precipitator of heavy metals. However, because of their low reproduction rates, anaerobes require a long start-up time and are sensitive to toxic shocks.37 Both aerobic and anaerobic processes have been shown to be capable of degrading landfill leachate.38 However, many landfill leachate treatments have been found to be insufficient if the anaerobic process is used alone without the aerobic. Systems comprising combined anaerobic-aerobic treatment are therefore recommended to achieve effective treatment at landfill.
The rate of mineralization of organic carbon in a biological process depends on the concentration of active cell mass. The maximum cell growth in a process will depend on nutrient availability, gas transfer, and toxicity of the leachate. In aerobic and anaerobic treatment lagoons, no provision is made for concentrating the suspended cells. Therefore, lagoons must have a large surface area to facilitate effective organic destruction. The advantage of lagoons is that very low maintenance is needed except for a periodic desludging of the microbial sludge.20
The reduction in organic carbon achievable by microorganisms is limited to some extent by the minimum concentration required to maintain cellular metabolic processes.39 Microbial species known as oligotrophs can operate at low substrate concentrations, but they may not be able to reduce contaminant concentrations below water quality standards. There are methods to circumvent the biological maintenance barrier to leachate degradation. A well-known approach involves the use of activated carbon to enhance the biodegradation reaction.40 There are three known beneficial effects of adsorbent addition: organic carbon is concentrated for microbial attack in the microenvironment around the adsorbent particle; the concentration of potentially inhibitory organic compounds in the bulk solution is lowered; and the carbon particles serve as a surface for microbial growth.41
Leachate can also be degraded biologically in situ at the landfill site. Conditions within the landfill are controlled to encourage microbial activity, and leachate is recirculated through the landfill. Recirculated leachate may require nutrient amendment, neutralization, or heavy metal removal. Aerobic microbial activity occurs at the landfill surface, and anaerobic activity occurs in the landfill interior. Recirculation, combined with anaerobic activity, may stabilize heavy metals through the precipitation of heavy metal sulfides.42 Aerobic biodegradation is faster and better understood, and methods for encouraging aerobic activity within a landfill by the addition of hydrogen peroxide or air microbubbles have been investigated.43 Subsurface aeration wells have also been used to encourage in situ degradation.
Biodegradation is considered the first option for the primary removal of organic compounds from landfill leachate. However, some organic compounds are resistant to biological attack. In addition, biological sludge resulting from biological processes may become a disposal problem, particularly because of its capacity to store adsorbed undegraded hydrophobic organic species and heavy metals. No biological leachate treatment processes have yet to take advantage of microbial transformations, nor has adsorption of heavy metals though suitable microorganisms been studied in the laboratory.44,48 Bioremediation processes are still relatively unsophisticated and the potential exists for combining various types of microbial process schemes for selective component removal.9
15.8.1 Case 1: Anaerobic/Aerobic Treatment of Municipal Landfill Leachate in Sequential Two-Stage Up-Flow Anaerobic Sludge Blanket Reactor (UASB)/Aerobic Completely Stirred Tank Reactor (CSTR) Systems
A project was conducted to study the treatability of leachate produced from a laboratory-scale simulated reactor treating food wastes using a two-stage sequential up-flow anaerobic sludge blanket reactor (UASB)/aerobic completely stirred tank reactor (CSTR).45 Experiments were performed in two UASB reactors and a CSTR reactor having effective volumes of 2.5 and 9 L, respectively. The hydraulic retention times in the anaerobic and aerobic stages were 1.25 and 4.5 d, respectively. Following the startup period, the COD concentration of the leachate steadily increased from 5400 to 20,000 mg/L. The organic loading rate (OLR) was increased from 4.3 to 16 kg/m3/d by increasing the COD concentrations from 5400 to 20,000 mg/L.
As reported, the effluent of the first anaerobic UASB reactor (Run1) was used as the influent of the second UASB reactor (Run2), and the effluent of the second UASB reactor was used for the influent of the aerobic CSTR reactor (Run3). COD removal efficiencies for the first UASB reactor and in the whole system (two-step UASB + CSTR) were 58%, 62%, 65%, 72%, 74%, 79%, and 96%, 96.8%, 97.3%, 98%, 98%, and 98%, respectively. As the OLR increased from 4.3 to 16kg/m3/d, the COD removal efficiency reached a maximum of 80%. NH4-N removal efficiency was ca. 99.6% after the aerobic stage. The maximum methane percentages of the first and second UASB reactors were 64% and 43%, respectively.
The study used two continuously fed stainless steel anaerobic UASB (2.5 L) reactors and an aerobic CSTR reactor (9 L). The UASB was operated at 37 to 42°C using an electronic heater located in the central part of the reactor. The system was provided with a settling compartment (with an effective volume of 1.32 L). The dissolved oxygen concentration was maintained above 2 mg/L in the CSTR reactor. Partially granulated anaerobic sludge taken from the methanogenic reactor of the Pakmaya Yeast Baker Factory in Izmir was used as seed in the UASB reactor. The activated sludge culture was obtained from the DYO Dye Industry in Izmir and was used as seed for the aerobic CSTR reactor.
In this study, anaerobic and aerobic processes using sequential two-step UASB/CSTR reactors were found to form a feasible process for treating the leachate from food solid waste. COD removal efficiencies for the first and second anaerobic, aerobic and total system processes were 79%, 42%, 89%, and 98%, respectively. The COD loading rate used ranged from 4.3 to 16kg/m3/d.
The methane content of the first UASB reactor was ca. 60%. The NH4-N removal efficiency of the total system was 99.6%. Ammonium nitrogen was converted to nitrate in the aerobic system via nitrification. The BOD5/COD value obtained at the final stage was in the range 0.12 to 0.15.
15.8.2 Case 2: Comparison of Two Biological Treatment Processes Using Attached-Growth Biomass for Sanitary Landfill Leachate Treatment
Two biological systems were compared using attached-growth biomass for the treatment of leachate generated from a municipal waste sanitary landfill. A moving-bed biofilm process, which is a relatively new type of biological treatment system, was used.46 The process was based on the use of small, free-floating polymeric (polyurethane) elements, and biomass was grown and attached as biofilm on the surface of these porous carriers. For comparison, a granular activated carbon (GAC) moving-bed biofilm process was also tested. This method offered the advantages of combining both physico-chemical and biological removal mechanisms for the removal of pollutants. The presence of GAC in the reaction tank provided a porous surface able to adsorb both organic matter and ammonia, as well as to provide an appropriate surface onto which biomass could grow. A laboratory-scale sequencing batch reactor (SBR) was used for examination of both carriers. The effects of different operation strategies on the efficiency of these biological treatment processes were studied in order to optimize their performance, especially for the removal of nitrogen compounds and biodegradable organic matter. It was found that these processes were able to remove nitrogen content almost completely, and the removal of organic matter such as BOD5 and COD was acceptable.
The SBR reactor used was constructed from cylindrical Plexiglas® with a working capacity of 8 L (as shown in Figure 15.1).47 The contents of the reactor were mixed with a magnetic stirrer, and
Leachate flow in
Leachate flow in
Top water level
Air bubbles wm/i r
FIGURE 15.1 Schematic diagram of a laboratory-scale, sequencing batch (bio) reactor (SBR).
a ceramic diffuser was used for aeration. A peristaltic pump was used to feed leachate directly into the SBR, as well as to remove the treated effluent.
The study consisted of two separate treatment cycles using a suspended-carrier attached-biofilm process. During the first cycle, the SBR was filled up to 50% of its empty volume with cube-shaped waste polyurethane particles (total dry weight 30 g). The density of the carrier media was slightly below 1 g/cm3, so the waste particles could easily follow the water flow pattern, circulating in the filled reactor. The continuous motion eliminated problems with clogging and dead space, which can often decrease the efficiency of fixed-bed biofilm systems. The cubes (having an approximate dimension of 1 cm) present high porosity (20 to 40 pores/cm2). During the second operational cycle, GAC (type F400, supplied by Chemviron Co., Belgium) was added to the reactor (90 g total), with a specific surface area of 1100 m2/g and density of 1.2 g/cm3. The main parameters studied during this investigation included the following:
1. The addition of alkalinity, phosphorus, and methanol (different concentrations and rates were evaluated)
2. An increase in the hydraulic retention time
3. A replacement sequence for used carrier media
4. The application of intermittent aeration, i.e., operation with alternate aerobic and anoxic conditions
Table 15.1 summarizes process efficiency during the first operation cycle of the SBR, and Table 15.2 shows the treatment results for the second operation cycle (GAC).
This study demonstrated that the suspended carrier-biofilm treatment method can offer an alternative option to the conventional activated sludge process for the effective removal of carbon and nitrogen in sanitary landfill leachates. Although raw leachate is very difficult to treat, complete nitrification and a high degree of organic carbon removal were achieved using the moving-bed biofilm SBR process.
The study reported some problems regarding the data for the biofilm from the media after 3 weeks of operation, and also sludge accumulation at the bottom of the bioreactor. It was also found that an external carbon source, such as methanol, was necessary for controlling the denitrification stage.
An alternative moving-bed biofilm SBR process using GAC has also been proven to be an effective treatment method for the removal of nitrogen from landfill leachates. This method can remove biodegradable organic carbon (BOD5) and COD. However, the main disadvantage of this process is the buildup of a large amount of residual suspended solids, hence increasing sludge disposal costs. An overall comparison between the two attached biomass biological treatment processes showed an advantage for the process that used porous polyurethane as its carrier material.46
Average Treatment Results during First Operation Cycle (SBR)
Total Removal (%)
Alkalinity (mg/L CaCO3)
SBR, sequencing batch reactor; TDS, total dissolved solids.
Treatment Results for the Second Operation Cycle (GAC)
BOD5, 5d-biochemical oxygen demand; GAC, granular activated carbon; TDS, total dissolved solids; NTU, normal turbidity units.
BOD5, 5d-biochemical oxygen demand; GAC, granular activated carbon; TDS, total dissolved solids; NTU, normal turbidity units.
15.8.3 Case 3: Leachate Treatment Using an Aerobic Biofilm Reactor
In this project, leachate was treated using an innovative aerobic biofilter utilizing special plastic media. Aerobic biofilters have been shown to be very effective in many treatments for removing organic pollutants and also their nutrient content. This study focused on leachate treatment using an attached growth biofilm reactor, which contains a packing of 80 mm diameter plastic media called "Cosmo balls."47 Figure 15.2 shows how the experiment was set up. The selected parameters for the study include COD, ammonia nitrogen, pH, and BOD. The results showed that the COD removal percentages were above 90% for COD but declined to 70% at very high loading. The ammonia nitrogen removal achieved in the study was above 85%.
The use of an attached growth aerobic biofilm reactor to treat leachate is relatively new. Past studies on anaerobic biofilters showed excellent organic removal up to 90%, and the retention time needed to treat high-strength effluent was between 3 and 5 d. The use of aerobic biofilters using
FIGURE 15.2 Schematic diagram of leachate treatment using an attached growth biofilm reactor.
FIGURE 15.2 Schematic diagram of leachate treatment using an attached growth biofilm reactor.
122 136 153 173 196 233 238
FIGURE 15.3 Relationship of COD content (mg/L) of influent and effluent over time (d).
Cosmo balls has been successful in treating sewage effluent with a short hydraulic retention time of only 4 h. This study was carried out to evaluate process efficiency using Cosmo balls with a hydraulic retention time of 5 d.
The aerobic biofilter used in the study had a capacity of 10 L. The reactor was packed to 60% of the empty bed volume with Cosmo ball media. The biofilter was seeded with active innoculum taken from an active aerated lagoon of a nearby landfill leachate treatment. Fresh raw leachate was used as feed to the reactor at a rate of 5 L/d over 24 h. The loading rates applied to the bioreactor were between 1.6 and 22.2 kg COD/m3 d. Initial studies were conducted as a batch process lasting for a period of 24 d. Thereafter, the biofilter was fed continuously for a total period of 240 d.
Figure 15.3 shows that percentages of COD removed in the biofilter decreased with increasing feed COD concentration. The value of the influent fluctuated, indicating that leachate characteristics were never uniform. The aerobic bioreactor was shown to be capable of treating leachate with about 80% COD removal using the designed hydraulic retention time of 5 d. Figure 15.4 shows that the ammonia nitrogen levels in the treated effluent were fluctuating and that the percentage of ammonia nitrogen removed declined very slightly at increased ammonia loading. Ammonia nitrogen removal showed very good results, with more than 80% destruction achieved in this study.
122 136 153 173 196 233 238
FIGURE 15.3 Relationship of COD content (mg/L) of influent and effluent over time (d).
It was observed that the factors contributing to the variation of leachate data are solid waste characteristics, for example, the composition and size of the waste and degree of compaction, the moisture content and degree of rainwater infiltration, temperature, sampling, and analytical methods.47
BOD Bochemical oxygen demand
BOD5 Five-day biochemical oxygen demand
COD Chemical oxygen demand
CSTR Completely stirred tank reactor
DDT Dichloro diphenyl trichloroethane
DNA Deoxyribonucleic acid
GAC Granular activated carbon
MBAS Methyl blue active substances
MMO Methane mono-oxygenase enzyme
MSW Municipal solid waste
OLR Organic loading rate
PCPA Pentachlorophenol-induced periplasmic protein
SBR Sequencing batch reactor
TOC Total organic carbon
TDS Total dissolved solids
UASB Up-flow anaerobic sludge blanket reactor
VDS Volatile dissolved solids
VSS Volatile suspended solids
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