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Hydrochemistry and natural attenuation in leachate from tropical landfills

Mangimbulude, J.C.

2013

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Download date: 19. Apr. 2022

Hydrochemistry and natural attenuation

in leachate from tropical landfills

Cover: Janine Mariën Layout: Désirée Hoonhout Printing: VU University Press Amsterdam Thesis 2013-01 of the Department of Ecological Science, VU University Amsterdam, the Netherlands

VRIJE UNIVERSITEIT

Hydrochemistry and natural attenuation in leachate from tropical landfills

ACADEMISCH PROEFSCHRIFT

ter verkrijging van de graad Doctor aan de Vrije Universiteit Amsterdam, op gezag van de rector magnificus

prof.dr. L.M. Bouter, in het openbaar te verdedigen

ten overstaan van de promotiecommissie van de Faculteit der Aard- en Levenswetenschappen

op maandag 6 mei 2013 om 15.45 uur in de aula van de universiteit,

De Boelelaan 1105

door

Jubhar Christian Mangimbulude geboren te Tobelo, Indonesië

promotor: prof.dr. N.M. van Straalen copromotor: dr. W.F.M. Röling

Contents

1. General Introduction 7 2. Seasonal dynamics in leachate hydrochemistry and natural

attenuation in surface run-off water from a tropical landfill. 35 3. Microbial nitrogen transformation potential in surface run-off

leachate from a tropical landfill. 61 4. Microbial community structure and dynamics in relation to

hydrochemistry of surface run-off water from the Jatibarang landfill in Indonesia 83

5. Hydrochemical characterization of a tropical, coastal aquifer

affected by landfill leachate and seawater intrusion. 101 6. General Discussion 117 Samenvatting 125 Summary 129 Ringkasan 133 Acknowledgements 137

7

1 GENERAL INTRODUCTION

1.1. An overview of world-wide solid waste management All over the world, generation of municipal solid waste (see Box 1.1) has grown tremendously over the last few decades (Fig. 1). Although there is a wide variation among countries, on a per capita basis, waste generation in Europe is significantly lower than in the USA. In 2006, the Western European countries produced about 550 kg of municipal solid waste per person annually compared to 764 kg in the US (European Environment Agency, 2007). Overall waste generation rates in the European countries and the USA are, however, higher than those in countries of the Association of South-East Asian Nations (ASEAN) (Fig. 2). This is associated with a higher standard of living, higher consumption, and industrial developments (Daskalopoulos et al., 1998; UNICEF, 2001; Visvanathan and Glawe, 2006). It is impossible to eliminate solid waste production completely because it is a consequence of human activity. What can be done is to manage the waste streams properly. However, many cities in ASEAN countries have difficulty managing waste due to institutional, financial, technical, regulatory, knowledge, and public participation shortcomings. The consequence is environmental degradation and a serious risk to human health due to inadequate disposal of waste (Barlaz et al., 2003; Goorah et al., 2009; Ngoc and Schnitzer, 2009). In European countries, specific policies on solid waste management have been developed and these policies involve complex and multifaceted trade-offs among a range of technological alternatives, economic instruments, and regulatory frameworks. For instance, there are several waste management options after generation and before final disposal: (a) waste minimisation, (b) collecting and sorting, (c) re-use, (d) recycling, (e) composting, (f) incineration and (g) dumping in a landfill (Strange, 2002) (see Box 1.2). The European member states show significant differences with respect to standards and practices of solid waste management. The EU recommends that member countries develop an integrated approach to waste management in order to contribute to sustainable urban development in Europe (Committee on the Environment, Agriculture and Local and Regional Affairs, 2005). Waste management practices have various environmental, economic, social, and regulatory implications not only for local policies, but also for global sustainable development (Pires et al., 2011). Integrated solid waste management should be environmentally effective, socially acceptable for a particular region, and should provide opportunities for local economies and improvement of the quality of life. Nowadays, municipal solid waste management has become an issue of routine policy in developed countries. In principle it is an integrated management approach under the responsibility of local authorities. This approach includes all options of waste generation, reduction, re-use, recycling, storage, sorting and collecting, transport and final disposal or incineration.

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8

Box. 1.1. Definitions of solid waste The term municipal solid waste (MSW), refers to solid wastes from houses, streets and public places, shops, offices, and hospitals; collection and disposal of such waste is often the responsibility of municipal or other governmental authorities (Zurbrugg, 2003). Solid wastes are all materials arising from human activities, that are normally solid and are discarded as unwanted. Because of their immanent properties, discarded waste materials are often reusable and may be considered a resource in another setting. Solid waste management is the term applied to all of the activities associated with the management of society’s waste (http://assembly.coe.int/Main.asp?link=/Documents/WorkingDocs/Doc07/EDOC11173.htm). MSW management encompasses planning, engineering, organization, administration, financial, and legal aspects of activities associated with generation, growth, storage, collection, transport, processing and disposal in an environmentally compatible manner, adopting principles of economy, aesthetics, energy and conservation (Tchobanoglous et al., 1993). MSW includes commercial, non-hazardous industrial waste and potentially also demolition waste and sewage sludge. Depending on the country, the definition can include some or all of:

• Regular household waste • Household hazardous waste • Bulky waste from households • Street sweepings and litter • Park and garden waste • Waste from institutions, commercial establishments and

offices • Construction and demolition waste • Sewage sludge

From 1950 to 2000 the world’s economic activities increased fifteen-fold. The growth of consumer societies all over the world has seen a large increase in solid waste produced per head, and the waste mix has also become ever more complex. For more than 50 years we have taken for granted that our waste could be deposited in holes in the ground or incinerated. In the urbanized world cities use the bulk of the world’s resources and discharge most waste (Waste and Sustainable Cities, 2007).

General Introduction

9

Fig. 1. Annual rates of solid waste generation in the USA (left) and selected European countries (right); Denmark (♦), The Netherlands (◊), France (▲) and Sweden (∆). (Sources: European Environment Agency, 2007; Environmental Protection Agency, 2007).

Fig. 2. Rates of waste generation in the ASEAN nations, USA and European countries (Sources: Visvanathan and Glawe, 2006; Chaerul et al., 2007; European Environment Agency, 2007; Environmental Protection Agency, 2007). 1.2. Municipal solid waste management in practice in Indonesia Indonesia is a developing country with a total land area of 1.9 billion km2 hosting a population of 237 million (2010). In the densely populated areas, disposal of solid waste is a major environmental problem and has become a national issue. The average per capita waste generation in Indonesia is 0.76 kg/day. Thus, with all inhabitants jointly, Indonesia is estimated to generate about 187,960 ton/day of municipal waste (Chaerul et al., 2007). Municipal solid waste composition in Indonesia is dominated by organic matter; it stems mainly from household sources (43.4%) (Meidiana and Gamse, 2010). Basically, waste management in Indonesia includes generation and storage systems, collecting systems,

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USAEuropean

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Waste generation rates (kg/cap/day)

Chapter 1

10

transportation systems and final disposal systems. According to the Japanese International Corporate Agency (2008) waste management in Java is able to serve 59% of the total population. It was reported that the level of waste management services at the national level only reaches 56%, while the remaining is handled by civilians. Padmi (2006) stated that waste not managed by the government is handled locally using burning (35%), burying (7.5%), composting (1.6%), or other means (15.9%). This situation is still valid nowadays, including the city of Surabaya. Overall, solid waste fluxes and pathways in Indonesia can summarized as indicated in Fig. 3.

Fig. 3. Schematic diagram of municipal solid waste management in Indonesia (Chaerul et al., 2007). Municipal solid waste management in developing countries faces a number of specific problems. These can be described as (Zurbrugg, 2003): 1) inadequate coverage of municipal services, and operational inefficiency of services, 2) limited utilization of recycling activities, 3) inadequate landfill disposal, and 4) inadequate management of hazardous and health-care waste. In fact, typical MSW management in developing countries shows an array of problems, including low collection coverage and irregular collection services, crude open dumping, lack of water pollution control and the handling and limited control of informal waste picking or scavenging activities (Ogawa, 2000). As a developing country, urbanization in Indonesia is an ongoing process, and the management of solid waste is becoming a major environmental and public health problem in urban areas. These problems are caused by technical, financial, institutional, economic, and social factors, which constrain the development of effective solid

General Introduction

11

waste management systems (Ogawa, 2000). Problems occur in all steps of waste management: storage, collection, transporting, and treatment, with a tendency to aggravate towards the end of the chain, the landfill. Therefore, a higher effort is needed including national legislation on solid waste management. Basically, until 2007 there was no national waste policy describing the concepts, aims and measures. A specific Waste Management Regulation has been issued recently (2008) by the Government of Indonesia. The regulation stated that the purpose of solid waste management is to improve public health and environmental quality and turn waste into a resource. Both the central and local governments should control adequate waste management in an environmentally friendly manner, in accordance with the objectives of the Regulation. To that purpose, the government cooperates with other parties including universities to develop and implement proper waste management including landfill leachate management. The 2008 Regulation can be an important legal tool in forcing all parties involved to support the national waste management strategy (Meidiana and Gamse, 2010). 1.3. Waste management: social aspects The social component of solid waste management focuses on the interaction between local communities and the institutional arrangements for managing solid waste (LOGO South Thematic Programme, 2006). Areas of attention which are important in this respect are community awareness, orientation of the municipal service to the real service needs of the population, to mobilise and support community groups to take a role in the solid waste management in their area (community participation). A study published by Saribanon et al. (2009) concluded that increasing public awareness and changing the public’s perception of solid waste are key to the social aspects of solid waste management. Increased public awareness will increase community participation in waste collection activities and enhance local separation and temporary storage of waste in the households. Saribanon et al. (2009) also reported that community-based management is an important strategy for solid waste management in Indonesia. Increasing participation through community-based management is more effective than changing people’s perception and behavior. For instance, in some cases, a change in the behavior of a community towards improved separation and recycling took more than 10 years while with appropriate support, community-based management only took 2 years. Community participation in solid waste management has become important since many local authorities in developing countries suffer from a lack of financial, technical and human resources for basic urban services (Desai, 1994; Moningka, 2000). Community participation in solid waste management is required particularly for solid waste collection, separation, and delivery in human settlements. In contrast to developed countries, in developing countries such as Indonesia, the Philippines, Nigeria, Vietnam, Bangladesh, and India, there are people that make a living on landfills. They collect items from the garbage and recycle waste materials, which are then sold. These people are called waste pickers or scavengers (Johannessen and Boyer, 1999). The scavengers risk their health in the process of recycling waste materials from a landfill (see Box 1.3). The Indonesian government officially does not allow scavengers to live and work on a landfill for public health risk reasons, but still scavenging continues because these people often have no other means to make a living, while the government is not able to provide

Chapter 1

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employment opportunities for them. Besides, most people will not accept the presence of scavengers within residential areas, so they are bound to practice scavenging on the landfill. One should realize that the scavengers contribute a lot to the process of re-using and recycling waste materials (Suwartiningsih, 2010). To ensure their health, without removing their source of revenue, the government only allows scavengers to work as waste pickers in the street or in shelters for temporary garbage. By this way the authorities aim to improve the health status of the scavengers.

Box 1.2. The Solid Waste Management Hierarchy The Waste Management Hierarchy is an internationally recognised strategy for management of municipal solid wastes. It places greatest emphasis on strategies and programs for avoiding and reducing waste, with treatment and disposal being the least favoured options (United Nations Environment Programme (UNEP), 2005).

Hierarchy from highest to lowest priority:

Reduce

Re-use

Recycling

Incineration

Landfilling

Increasing cost

General Introduction

13

1.4. Chemistry of landfill leachate Landfills produce leachate that may mix with the groundwater, drainage systems or surface run-off. The generation of such leachates and their chemical composition varies widely and is significantly influenced by the type of waste, seasonal precipitation, moisture, temperature, landfill geometry and the age of the waste (Chu et al., 1994; Kjeldsen et al., 2002; Renou et al., 2008).

Box 1.3. Scavenger communities Suwartiningsih (2010) has studied the scavenger community of the Jatibarang landfill, near Semarang, Central Java. The author reported that there are around 700 scavengers living on the site today. They come from several towns around Semarang like Purwodadi, Bawen, Karanggede and Boyolali. There are two types of scavengers who settle in Jatibarang. First, the scavengers who depend only on collecting garbage, second, the scavengers who collect garbage after planting or harvesting crops in their hometown. In depth interviews conducted by Suwartiningsih revealed that there are several reasons to become a scavenger. First, there is no other choice, second, mimicking a successful brother or neighbor of the activities of collecting garbage and third, waiting for harvest time. Garbage scavengers are a portrait of poverty in developing countries. They must struggle and take risks in a marginal position in society if they are to survive. They usually have a low education, most of them finished only elementary school, and some have not even graduated from elementary school. However, scavengers have an important role in the process of collecting and recycling waste. They collect garbage for on average 10 hours per day, while the amount of useful items collected reaches 50 - 100 kg per person. This means that each scavenger reduces landfill waste by as much as 1500 - 3000 kg per month. With a total scavenger population of 700 persons, it means that waste reduction on the Jatibarang landfill through scavenging activities may reach 1050 to 2,100 tons per month. The average income of scavengers in Jatibarang reaches Rp 50,000 per day per person. This amount is higher than the income of a construction worker who only earns Rp 30,000 per day. Perhaps this is also the reason for people to become a scavenger,

Chapter 1

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Typically, leachates can be categorized into four major groups: (a) dissolved and particulate organic matter (b) inorganic materials and nutrients (c) heavy metals, and (d) hazardous compounds such as aromatics and halogenated compounds. There is no single specific parameter to characterize landfill leachate; some common parameters often used to classify leachates are given in Table 1. Table 1. List of physico-chemical parameters by which landfill leachates are usually characterized. Category Parameters General parameters pH, temperature, total suspended solids, electrical conductivity, alkalinity, dissolved oxygen, total dissolved N Organic matter BOD (Biochemical Oxygen Demand) COD (Chemical Oxygen Demand) TOC (Total Organic Carbon) and DOC (Dissolved Organic

Carbon) Inorganic matter and heavy Cations: Fe, Mn, Ca, K, Mg, Na, NH4 metals Anions: Cl, HCO3, SO4, NO3, NO2 Trace metals: Cu, Hg, Zn, Cd, Pb, Ni Organic compounds phenol, benzene, toluene, xylene, volatile organic compounds,

chlorinated aliphatics, chlorinated aromatics, polycyclic aromatic hydrocarbons, pesticides, etc.

Sources: Jensen et al. (1999); Christensen et al. (2001); Kljeldsen et al. (2002); Öman and Junestedt (2008). Many landfill leachates contain high concentrations of pollutants. This includes ammonium and aromatic compounds such as phenols (Burton and Watson-Craik, 1998). These compounds are often found in concentrations above the standards for discharge in water. Three types of leachate can be defined, depending on landfill age (Table 2). A young leachate in the acidogenic phase is characterized by a high organic fraction and a BOD/COD ratio greater than 0.4. Older leachate, in the methanogenic phase, is not as easily biodegraded as a young leachate. It contains refractory organic compounds, is usually highly contaminated with ammonia resulting from hydrolysis and fermentation of nitrogen-containing fractions of biodegradable refuse (Cheung et al., 1997) and is characterized by high pH values. The influence of tropical climatic conditions on changes in leachate quality and quantity should be considered when designing a landfill for disposal of waste and proposing a treatment scheme for the leachate (Asian Regional Research Program on Environmental Technology (ARRPET), 2004; Tränkler et al., 2005). In hot humid climates, leachate quantity is much higher than in hot arid regions. On the other hand, high rainfall leads to reduced leachate strength due to dilution (Tränkler et al., 2005).

General Introduction

15

Table 2. Landfill leachate types according to age (Alvarez-Vazquez et al., 2004). Characteristics Young Medium Old Age (year) < 1 1 - 5 > 5 pH < 6.5 6.5 – 7.5 > 7.5 COD (mg L-1) > 15,000 3000 – 15,000 < 3000 BOD/COD 0.5 - 1 0.1 – 0.5 < 0.1 TOC/COD < 0.3 0.3 – 0.5 > 0.5 NH3-N (mg L-1) < 400 400 > 400 Heavy metals (mg L-1) > 2 < 2 < 2 Organic compounds 80% VFA 5-30% VFA +

HA + FA HA +FA

only VFA: volatile fatty acids, HA: humic acids, FA: fulvic acids Temperature exerts an influence on leachate quality by controlling the rate of bacterial activity. Microorganisms are able to grow in refuse over a temperature range of 5 – 55 oC, but the highest growth rate occurs between 15 – 30 ºC (Dobson, 1964). A drop of 5 oC results in a 10 - 20% decrease in the rate of bacteria-controlled oxidation of organic matter (Camp, 1963). Thus, in a temperate climate significantly less BOD is produced from a landfill during the winter and spring months than during summer and fall (Apgar and Langmuir, 1971). The majority of landfills in Indonesia are open-dump systems which lack liners or proper drainage systems to divert the leachate. This poses the problem that the leached material can reach soil, surface water and also groundwater. Groundwater contaminated with landfill leachate will migrate through the unsaturated zone following the hydraulic gradient of the groundwater system. The direction (i.e. vertical or horizontal) and velocity of leachate flows may differ physically from the groundwater depending on the local water table gradients, the viscosity and the density of the leachate. The water table gradient is dependent on the hydrogeology of the immediate area, the viscosity is a factor of the texture of the dissolved solids in the leachate and can cause reduced leachate-flow velocities and dilution rates and finally, the density of the leachate relies on the temperature and concentration of dissolved solids (Bjerg et al., 2003). Variations between leachate and groundwater densities can cause considerable discrepancies in the vertical positioning of the leachate plume below the landfill (Bjerg et al., 2003). 1.5. Microbiology of landfill leachate Microbial respiration is tightly coupled to the redox chemistry of landfill leachate or the groundwater (Lovley and Phillips, 1998; Baun et al., 2003). Redox components are important factors for microbial metabolism, because they can be used by microorganisms as electron acceptors to degrade dissolved organic carbon in leachate through the aquifer and away from the landfill. The migration of dissolved constituents, such as metals and organic compounds, is substantially affected by redox components and redox reactions (Fig. 4). The order of oxidation of organic carbon by electron acceptors according to the Gibbs free energy at a neutral pH is: O2, NO3

-, Mn(IV), Fe(III), SO42- and CO2 (Van Breukelen et al., 2003).

According to Christensen et al. (2001), the overall sequence of conditions in the redox

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gradient of a leachate begins with methanogenic conditions close to the landfill, and runs through SO4

2-, Fe2+, Mn2+ and NO3- reducing conditions to aerobic conditions furthest away

from the landfill. These microbial processes typical for landfill leachate, such as denitrification, manganese reduction, iron reduction, sulfate reduction, and methane generation, are shown in Table 3. Some of these metabolic types (sulfate reducers and methanogens) can be subdivided into different physiological groups depending on the electron donor, as observed in the Norman Landfill (USA) leachate-contaminated aquifer (Beeman and Suflita, 1987, 1990; Harris et al., 1999).

Figure 4. Spatial distribution of redox processes and degradation of dissolved organic carbon (DOC) in a typical landfill leachate plume. Methane, Fe(II) and ammonium compete with DOC for electron acceptors at the plume fringe. Taken from Van Breukelen et al. (2003). Table 3. Most prominent redox process in landfill leachate plumes (Bjerg et al., 2003). Reaction Process 1. Methanogenic 2CH2O → CH4 + CO2 2. Sulfate reduction 2CH2O + SO4

2- + H+ → 2CO2 + HS- + 2H2O 3. Iron reduction CH2O + 4Fe(OH)3 + 8H+ → CO2 + 4Fe2+ + 11H2O 4. Manganese reduction CH2O + 2MnO2 + 4H+ → CO2 + 2Mn+ + 3H2O 5. Denitrification 5CH2O + 4NO3

- + 4H+ → CO2 + 2N2 + 7H2O 6. Aerobic respiration CH2O + O2 → CO2 + H2O 7. CO2 reduction HCO3

- + H+ + 4H2 → CH4 +3H2O 8. Nitrification NH4

+ + 2O2 → NO3- + 2H+ + H2O

9. Methane oxidation CH4 + 2O2 → HCO3- + H+ + H2O

General Introduction

17

In this thesis, microbiological studies covered two topics. First, an assessment of redox processes was conducted to evaluate the potential for natural attenuation of leachate and surface run-off. For this topic, different redox (aerobic and anaerobic) batch experiments were conducted to determine the capacities of leachate microorganisms to degrade organic matter (COD) and to transform nitrogen compounds. Second, most of the natural attenuation process of leachate contaminants took place through microbiological processes, so it is important to know the microbial community structure in leachate in relation to pollution and seasonal effects. The surface leachate microbial community structures described in this thesis were based on culture-independent molecular analysis of 16S ribosomal RNA genes isolated from surface leachate. Amplification of these markers by the polymerase chain reaction and subsequent fingerprinting by denaturating gradient gel electrophoresis (DGGE) provided microbial community profiles. The molecular analysis approach used in this research was developed in studies of subsurface leachate community structure at the Banisveld landfill by Röling et al. (2001) and Lin et al. (2005). This approach produces good data for microbial community profiles. Until now only little information is available about microbial community structures of surface leachates compared to subsurface or leachate plume communities. Inside a leachate plume the environment is characterized by the presence of reduced chemical species and high concentrations of dissolved organic matter. This environment is partly due to the composition of the leachate from the landfill and partly due to microbial processes in the plume (Bjerg et al., 1995). Microbial communities in landfill leachates were studied by, among others, Christensen et al. (2001), Ludvigsen et al. (1999), Röling et al. (2000), Albrechtsen et al. (1995), and Lin et al. (2005). The microbial community in landfill leachate is dominated by bacteria, as shown by analysis of phospholipid fatty acid profiles (PLFAs) (Ludvigsen et al., 1999). The bacterial numbers determined by direct counting range from 0.04 – 1470 x 106 cells g-1 dry weight (Christensen et al., 2001; Albrechtsen et al., 1995). Because the structure and function of microbial communities exhibit significant spatial and temporal variability, collecting microbial data should be complemented by hydrochemical analyses of leachates. This combination of microbial and hydrochemical data is essential in evaluating and implementing remediation strategies. A study of landfill leachate or groundwater microorganisms can be done using the classical approach, involving culturing of microorganisms in solid or liquid growth media containing appropriate carbon and electron acceptor sources, and a range of other physiological conditions to promote microbial growth (Dos Santos et al., 2009). However, general culture conditions impose a selective pressure, preventing the growth of many “uncultivable” microorganisms. Some studies suggest that only a small fraction (1 - 15%) of microbial species are cultivable under laboratory conditions and more than 85% have never been studied, which means that culturing techniques provide only a narrow view of the actual microbial community (Amann et al., 1995; Pace, 1997). The use of molecular tools that do not depend on cultivation has expanded considerably our knowledge of landfill leachate and groundwater microbial diversity. For instance molecular biology techniques most often applied to wastewaters and leachates include: denaturing gradient gel electrophoresis (DGGE), terminal restriction fragment length polymorphism (T-RFLP), fluorescent in situ hybridization (FISH), and cloning of 16S rDNA (Sanz and Kochling, 2007). Recent studies show that microbial community structure in the water phase of a landfill leachate plume is clearly different from the community structure outside the plume, as shown

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in the Banisveld aquifer in the Netherlands by 16S ribosomal DNA-based denaturing gradient gel electrophoresis (Röling et al., 2001; Lin et al., 2005). Röling et al. (2001) found that there is a relationship between the dominant redox processes and the bacteria identified. In the iron-reducing plume, members of the family Geobacteraceae (Deltaproteobacteria) made a strong contribution to the microbial communities. Members of the β-subdivision of the phylum Proteobacteria dominated upstream of the landfill, but this group was not encountered beneath the landfill where Gram-positive bacteria dominated. Further downstream, where the effect of contamination decreased, the community structure shifted partly back; the contribution of the Gram-positive bacteria decreased and the β-Proteobacteria reappeared (Röling et al., 2001). 1.6. Natural attenuation: nitrogen transformations and phenol degradation in leachate Landfill leachate usually contains high nitrogen concentrations in the form of organic nitrogen and inorganic nitrogen. These are products of deamination of proteins in the waste. Among the various forms of mineral nitrogen (ammonium, nitrate, nitrite), ammonium is usually the highest (Berge et al., 2005; Kjeldsen et al., 2002). Numerous studied reported that ammonium is the primary cause of acute toxicity of municipal landfill leachate (Kim et al., 2006; Clément and Merlin, 1995; Ernst et al., 1994). Also, high concentrations of ammonium could promote eutrophication in lakes and rivers (Jokela et al., 2002). Besides ammonia, phenolic compounds released into the environment are of high concern because of their potential toxicity. These compounds found in the leachate include phenol, cresols, chlorinated phenols and other substituted congeners of these compounds (Varank et al., 2011). Many substituted phenols, including chlorophenols, nitrophenols, and cresols have been designated as priority toxic pollutants by the U.S. Environmental Protection Agency (Boopathy, 1997; Pohland et al., 1998). There is a great concern about phenolic compounds from the viewpoint of environmental protection (Chakarova-Käck et al., 2006) because the intermediate metabolites resulting from degradation of these compounds are no less toxic than the parent compounds (Slack et al., 2005; Kolvenbach et al., 2007) and these compounds significantly inhibited nitrification (Kim et al., 2008). Overall, these two groups of components rank highest for their potential to pollute surface water and groundwater due to their toxicity. Their concentrations in waste water should be reduced as much as possible to achieve a safe level for discharge into the environment. Among the various environmental remediation methods, the process of natural attenuation has received quite some attention in last few decades. The basic principle of natural attenuation is that natural processes in the subsurface can reduce pollutant loads to a level low enough such that the plume does not pose a substantial risk to the groundwater or the drainage area (Christensen et al., 1994). Natural attenuation comprises a wide range of processes including dilution and binding to sediment, however, microbial degradation is the most important factor since it is responsible for intrinsic bioremediation: it decreases the mass and toxicity of the contaminants (Röling and Van Verseveld, 2002). Natural attenuation of nitrogen in landfill leachate can be conducted by microbial transformations (biotransformation), including mineralisation (i.e. ammonification), nitrification, denitrification, anaerobic ammonia oxidation, nitrogen fixation and dissimilatory nitrate reduction to ammonium (Fig. 5) (Barlaz et al., 1990; Im et al., 2001; Welander et al., 1997; Etchebehere et al., 2002; Price et al., 2003; Kulikowska and Klimiuk, 2004). In the scheme of nitrogen conversions in landfill leachate, nitrogen fixation and

General Introduction

19

assimilative nitrate reduction are not important (Kadlec and Knight, 1996), and for that reason these processes were not considered in this research. The most relevant processes are discussed below, with reference to Fig. 5.

Fig. 5. Generalized scheme of nitrogen transformations (modified from Crumpton and Goldsborough, 1998; Brock, 2006). 1.6.1. Ammonification Organic nitrogen comprises a variety of nitrogen containing compounds including proteins, amino acids, amino sugars, urea, uric acid, purines and pyrimidines (Kadlec and Knight, 1996; Barlaz et al., 1990). The nitrogen content of municipal waste is usually around 1% on a wet-weight basis (Abeliovich, 1985; Bonin, 1995) and is due primarily to proteins contained in garden waste, food waste, and biosolids (Burton and Watson-Craik, 1998). The conversion of organic nitrogen to ammonia-nitrogen by aerobic and anaerobic heterotrophic bacteria (called ammonification), releases amino acids by hydrolysis of proteins, followed by deamination and respiration of the amino acids to carbon dioxide, ammonia, and volatile fatty acids. Ammonification rates are faster under aerobic conditions and decrease if the conditions switch from aerobic to anaerobic. 1.6.2. Nitrification Nitrification in landfill leachate has been studied by several researchers (Welander et al., 1997; Kulikowska et al., 2010; Wang et al., 2010; Yusof et al., 2010). Under aerobic conditions, ammonium is oxidized by microorganisms to nitrate, with nitrite as an intermediary product (Paredes et al., 2007). Two different groups of bacteria play a role in nitrification: ammonium oxidizers (e.g. Nitrosomonas) and nitrite oxidizers (e.g Nitrobacter).

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These bacteria are generally known as nitrifiers. In addition, recently knowledge revealed that nitrification is not only done by bacteria, but also by Archaea. Nitrosopumilus maritimus represents the first isolated nitrifierwithin the domain Archaea (Könneke et al., 2005; Wuchter et al., 2006). Nitrification does not occur optimally at dissolved oxygen levels less than 2 mg L-1 (Grady, 1999). A number of environmental factors influence nitrification, (e.g. Van Dongen et al., 2001). Nitrification also requires a long retention time (Mendoza-Espinosa and Stephenson, 1999), a low food to microorganism ratio (F:M), and adequate buffering (alkalinity) with a pH of 6.5 to 8 (Sedlak, 1991; Grady, 1999). Kim et al. (2008) reported that free cyanide above 0.2 mg L-1 seriously inhibited nitrification, but ferric cyanide below 100 mg L-1 did not. Phenol and p-cresol significantly inhibited nitrification above 200 mg L-1 and 100 mg L-1, respectively. 1.6.3. Denitrification When oxygen in landfill leachate is depleted, nitrate produced by nitrification can be converted to nitrogen gas by numerous heterotrophic bacteria in four steps, using the substrates contained in landfill leachate as electron donors (denitrification). These steps include the reduction of nitrate to nitrite, nitric oxide and nitrous oxide, and finally to nitrogen gas. According to Paredes et al. (2007) the denitrification process can be enhanced by adding readily biodegradable carbon sources such as acetic acid and methanol. Welander et al. (1998) have studied nitrate removal from municipal landfill leachate in a biofilm and found that the maximum denitrification rate with methanol as a carbon source was approximately 55 g N/m3/h. As the process had reached optimal operation, inorganic nitrogen was almost completely removed and the removal of total nitrogen was approximately 90%. Sometimes denitrification in landfill leachate is not complete due to nitrite accumulation. Nitrite is detrimental to many organisms so if wastewater contains high nitrite concentrations, this can be a big problem for the operation of activated sludge processes (Xuezheng et al., 2010). Nitrite accumulation was investigated in a sequencing batch reactor (SBR) by which a pre-treated landfill leachate was treated in an anoxic/anaerobic up-flow sludge bed (UASB) (Sun et al., 2009). According to Xuezheng et al. (2010) nitrite accumulation is caused by the difference between nitrate reduction rate and nitrite reduction rate. It also results from the absence of some species of bacteria or enzymes in the denitrification process. Cortez et al. (2001) have studied removal of nitrate from mature landfill leachate with a high nitrate load in a lab-scale anoxic rotating biological contactor. Under a phosphate concentration of 10 mg P-PO4

3− L−1 and nitrate concentrations above 530 mg N-NO3− L−1 the reactor achieved

nitrogen-nitrate removal efficiencies close to 100%, without nitrite or nitrous oxide accumulation. Ignoring nitrite accumulation, nitrate reduction is often used as the sole parameter to characterize denitrification, however, more accurately, the reduction of total oxidized nitrogen (nitrate + nitrite) should be used as a measure (Sun et al., 2009). 1.6.4. Anaerobic ammonium oxidation (anammox) The anammox process was first described in bioreactors of wastewater treatment plants in Delft, The Netherlands (Mulder et al., 1995). In this process ammonium reacts with nitrite as an electron acceptor to produce dinitrogen gas (Van de Graaf et al., 1996). Hydrazine (N2H4) as been identified as intermediate of the anammox process (e.g. Van de Graaf, 1997; Kartal et

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al., 2011). Further studies have shown that the anammox process is carried out by a group of bacteria, which to belong to the order Planctomycetales (Strous et al., 1999).

The main parameters with an inhibitory effect on anammox are dissolved oxygen (DO) while also organic compounds of low molecular weight have been suspected (Güven et al., 2004; Isaka et al., 2008). Mechanical stress was also discovered as an inhibitory factor (Arrojo et al., 2006). Carbon dioxide is used as the main carbon source for growth. However, besides growing chemo-autolithotrophically, anammox bacteria can also use organic acids as electron donors with nitrate as electron-acceptor (Güven et al. 2005; Kartal et al. (2007)). The oxidation of organic compounds coupled to nitrate reduction may contribute significantly to the success of anammox bacteria. Kartal et al. (2007) also reported that metabolism of organic sources in anammox bacteria (e.g K. stuttgartiensis) is similar to denitrification, they use nitrate directly and even out-compete heterotrophs for organic acids in the presence of ammonium, Overall nitrogen gas production by anammox bacteria through nitrate reduction proceeded via two reactions: dissimilatory nitrate reduction to ammonium followed by anaerobic oxidation of ammonium (Kartal et al., 2007). The anaerobic oxidation of ammonium to N2 has also been found in various natural environments (Dalsgaard et al., 2005). Kuypers et al. (2005) reported that anammox bacteria to be present and active in the oxygen minimum zone of the Benguela upwelling system. Ruscalleda et al. (2008) have applied the anammox process to treat urban landfill leachate coming from a previous partial nitrification process. In the presence of organic matter, the anammox process could coexist with heterotrophic denitrification. Nitrite accumulated in denitrification may serve as electron donor for the anammox reaction. 1.6.5. Nitrogen fixation Nitrogen fixation is the reduction of N2 from the atmosphere to ammonia by microorganisms. The microorganisms performing this process are known as diazotrophs and they provide about 60% of the annual nitrogen input into the biosphere (Newton, 2007). Diazotrophs are found in the domains Bacteria and Archaea and their lifestyle varies from free-living, loosely associated to symbiotic. Nitrogen-fixing microorganisms have also been found in ammonium-rich landfill leachate, but their physiological activities there are unknown (Wagner and Horn, 2006). 1.6.6. Dissimilatory nitrate reduction to ammonium (DNRA) DNRA occurs under anaerobic conditions; nitrate is used as an electron acceptor and is converted to ammonium (Price, 2003). DNRA is performed by fermentative bacteria (e.g., Enterobacteriaceae, Aeromonas, Vibrio spp., Clostridium sp.) (Philips et al., 2002). This process is influenced by nitrate and sulfide concentrations (Rütting et al., 2011; Tiedje, 1988). For example, under strongly reducing conditions where nitrate is limiting, DNRA has the advantage over denitrification since more electrons can be transferred per mole nitrate (Tiedje et al., 1982). Thus, DNRA is favoured when the electron acceptor is limiting relative to the electron donor (carbon), while denitrification is favoured when nitrate is higher (Smith 1982; Tiedje et al. 1982). Also, high sulfide concentrations induce the occurrence of DNRA (Vigneron et al., 2005). From an environmental viewpoint, DNRA in landfill leachate is not desired because it is expected to result in an increase in ammonium concentrations (Berge

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and Reinhart, 2005). However, a study conducted by Price et al (2003) showed that there was no significant increase in ammonium concentration due to DNRA. Biological systems have proven their potential for nitrogen removal, but the removal efficiencies are often unknown due to inadequate observation of nitrogen transformation and removal mechanism. Therefore is necessary to explore microbial nitrogen transformation in relation to modulating factors such as seasonal cycles. 1.6.7. Biodegradation of phenolic compounds Phenolic compounds may be defined as any compounds with aromatic nucleus bearing a hydroxyl group directly linked to the aromatic ring (Varadarajan, 2010). Based on the definition, phenolics include hydroxybenzoic acid, di- and trihydroxy-phenols, nitro-, chloro-, amino-, methoxy, phenoxy and alkyl-phenols (Varadarajan, 2010). Also included are some of the hydroxy derivatives of polyaromatic compounds (naphthols) and degradation products of pesticides (2,4-dichlorophenoxyacetic acid, 2,4,5-trichlorophenoxy acetic acid, 3-trifluoromethy-4-nitrophenol, and carbaryl) (Meulenberg, 2009; Varadarajan, 2010). Many phenolic compounds and pure phenols (Varadarajan, 2010) have been shown to be toxic to aquatic organisms (Gutes et al., 2005; Meulenberg (2009). Several substituted phenols, including chlorophenols, nitrophenols, and cresols, have been designated as priority pollutants by the US Environmental Protection Agency (Boyd et al., 1983). Phenolics and phthalates are two common hazardous organic compounds reported to occur in highly organic and heterogeneous municipal waste from urban sources in a tropical developing country (Swati et al., 2007). The amount of phenol in landfill leachate varies from one landfill leachate to another, but in general ranges from 0.6 to 1200 μg L-1 (Kjeldsen et al., 2002). In spite of the fact that phenolics are also toxic to many microorganisms (EPA, 1979), microbial biodegradation of phenolics has been demonstrated by many researchers (Boyd et al., 1983; Wang et al., 1988; Watanabe et al., 1998; Mardocco et al., 1999; Robles et al. 2000; Baek et al., 2001; Prieto et al., 2002; Zhang et al., 2004; Li et al., 2005; Bai et al., 2006; Chen et al., 2007; Tziotzios et al., 2007). Degradation of phenolic compounds may be carried out by eukaryotic and prokaryotic organisms under anaerobic and anaerobic conditions (Hughes et al., 1984, Fuchs, 2008; Sridevi at al., 2011). Under aerobic conditions oxygen is used as electron acceptor for energy-yielding (aerobic respiration) (Basha et al., 2010). Several studies reported that critical oxygen concentrations for bacterial cultures and yeast are in the range of 0.01 to 0.038 mg L-1 and 0.5 mg L-1 for flocculant microbial cultures (Longmuir, 1954; Gaudy and Gaudy, 1980). Besides that, dissolved oxygen is required as co-substrate in hydroxylation reactions (Fuchs, 2008). As oxygen is a co-substrate to activate phenolic compounds for further degradation under aerobic conditions, under anaerobic conditions different enzymes and degradation pathways need to be employed. Phenol degradation occurs via transformation to benzoyl-CoA, after wich the aromatic ring is reduced in the central benzoyl-CoA pathway (Fuchs, 2008). Some researchers have demonstrated that dissolved phenolic compounds can be degraded at relatively low concentrations by natural attenuation in shallow sand and gravel aquifers (Thornton et al., 2001).

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1.7. Study sites The sites used for the research presented in this thesis are the Jatibarang landfill in Semarang, Central Java and the Keputih landfill in Surabaya, East Java (Fig. 6).

1.7.1. Jatibarang landfill The Jatibarang landfill (Semarang, Central Java, Indonesia, coordinates 7o01’24.95” S, 110o21’32.23” E) has a surface area of 44.5 ha and is located in a hilly area at an altitude between 62 and 200 m above sea level, with slopes up to 25%, and 13 km from the city of Semarang. The landfill has been in operation as an open dump since 1992 until today (Fig. 6). It receives about 650–700 tons of waste per day from the city of Semarang. A local community of villagers lives close to the landfill and makes a living from scavenging and recycling valuable materials, as well as keeping cattle that feed on the waste. The leachate is collected in six connected, concrete ponds before entering the river Kreo.

Jatibarang Kèputih Fig. 6. Map of Java, showing the location of the two study sites, Jatibarang and Képutih. 1.7.2 Keputih landfill The Keputih landfill is located in the Eastern part of the greater Surabaya area, East Java, Indonesia (latitude 7o18.20’-7o19.20’and longitude 112o47.60-112o48.60). It has surface area of 24 hectares and is surrounded by traditional settlements (Fig. 6). The area where the landfill is situated was originally part of a natural salt marsh with a typical coastal wetland vegetation (e.g. mangrove and saltmarsh grasses). The area has an elevation of 3-4 m above sea level and the groundwater table is at 0 to 4 m below ground surface, depending on the season. The direction of regional groundwater flow is from West to East (Rachmansyah, 2001). The landfill received about 4000 m3 of solid waste daily including municipal waste, demolition waste, and industrial waste that is possibly hazardous (Endah and Pudjiastuti, 1995). The landfill was in operation since in the early 1980s and was closed in 2001. Part of the landfill is still used by scavengers and goats feed on the site.

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1.8. Objectives of the study In tropical countries seasonally varying climatic conditions should be taken into account when dealing with leachate or management of leachate. Most tropical countries have a distinct dry season of up to 150 days a year, and a wet season with intensive heavy showers and prolonged rainfall spells. Elevated temperature and high solar radiation influence both the quality and the quantity of landfill leachates (Tränkler et al., 2001). Taken jointly, these tropical climatic conditions may have a profound effect on the potential for natural attenuation of leachate pollutants. Against this background, the objectives of this thesis were: (1) to characterize seasonal changes of landfill leachates and their impact on groundwater

pollution surrounding the landfill, (2) to assess the potential for nitrogen tranformations and phenol degradation in landfill

leachate, (3) to profile seasonal changes of microbial communities present in landfill leachate and

groundwater and assess their contribution to natural attenuation. 1.9. Outline of the thesis Chapter 2 describes and discusses the seasonal dynamics of leachate hydrochemistry and the occurrence of natural attenuation at the Jatibarang site, Semarang. Monthly measurements of general landfill leachate parameters, organic matter-related factors and redox-related components were conducted. Field measurements in leachate ponds and laboratory experiments were done to reveal indicators of natural attenuation. Chapter 3 deals with microbial nitrogen transformations in Jatibarang leachate. Laboratory experiments were conducted to determine the seasonal influence on the initial rates of ammonification, nitrification, denitrification and anaerobic ammonium oxidation (anammox). Chapter 4 deals with a molecular analysis of microbial communities in the Jatibarang leachates. Diversity of microbial communities was characterized by means of denaturating gel electrophoresis applied to PCR-amplified 16S rRNA gene fragments and some functional genes indicative for aromatic biodegradation. Chapter 5 describes the hydrochemistry of groundwater contaminated by leachate in relation to seawater intrusion at the Keputih landfill. Chapter 6 summarizes the main results of this thesis and discusses their relevance for the process of natural attenuation of landfill leachate in tropical countries. References Abeliovich, A., 1985. Nitrification of ammonia in wastewater: Field observations and

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35

2 Seasonal dynamics in leachate hydrochemistry and natural attenuation in surface run-off water from a

tropical landfill

Mangimbulude, J.C., Van Breukelen, B.M., Krave, A.S., Van Straalen, N.M.,

Röling, W.F.M., 2009 – Waste Management 29, 829-838

Abstract Open waste dump systems are still widely used in Indonesia. The Jatibarang landfill receives 650-700 ton municipal waste per day from the city of Semarang, Central Java. Part of its leachate flows via several natural and collection ponds to a nearby river. Our aims were to identify seasonal landfill leachate characteristics in this surface water and to determine the occurrence of natural attenuation, in particular the potential for biodegradation, along the flow path. Monthly measurements of general landfill leachate parameters, organic matter-related factors and redox-related components revealed that leachate composition was influenced by seasonal precipitation. In the dry season, Electrical Conductivity and concentrations of BOD, COD, N-organic matter, ammonia, sulphate and calcium were significantly higher (1.1 to 2.3 fold) than during the wet season. Dilution was the major natural attenuation process acting on leachate. Heavy metals had the highest impact on river water quality. Between landfill and river, a fivefold dilution occurred during the dry season due to active springwater infiltration, while rainwater led to a twofold dilution in the wet season. Residence time of leachate in the surface leachate collecting system was less than 70 days. Field measurements and laboratory experiments showed that during this period hardly any biodegradation of organic matter and ammonia occurred (less than 25%). However, the potential for biodegradation of organic matter and ammonia was clearly revealed during 700 days of incubation of leachate in the laboratory (over 65%). If the residence time of leachate discharge can increased to allow for biodegradation processes and precipitation reactions, the polluting effects of leachate on the river can be diminished.

Key words: municipal solid waste, tropical landfill, leachate composition, natural attenuation

Chapter 2

36

1. Introduction Landfilling is widely applied as a disposal method for municipal solid waste (MSW) in developing countries. However, many landfills are not properly managed, and pose a serious threat to the environment due to leachate run-off. Landfill leachate can contaminate groundwater and surface water (Kjeldsen, 1993), rendering this water unsuitable for human use. In tropical countries, up to 90% of solid waste is disposed in open dump systems (Tränkler et al., 2005), and in many cases significant surface leachate run-off and gaseous emission are generated. The quantity and quality of leachate is influenced by the amount, composition and moisture content of the solid waste, as well as by local factors such as hydrogeological conditions, climate, height, and type of landfill (Johansen and Carlson, 1976; Chu et al., 1994). High local population densities, high annual precipitation (up 2000 mm), temperature (25 – 35 oC), and humidity (60%– 80%), as well as a distinct dry season of up to 150 days a year plus a wet season with frequent downpours, add to variation and complexity in landfill leachate characteristics and composition in tropical ecosystems (Miyajima et al., 1997). The composition of landfill leachates oftens shows significant seasonal variation (Christensen et al., 2001). Åkesson and Nilsson (1997) observed lower concentrations of leachate components during the wet autumn season for a Swedish landfill. Leachate treatment in tropical countries, in order to diminish environmental impacts, is a difficult task due to the seasonal flux of leachate pollutants (Chu et al., 1994; Chen 1996; Fan et al., 2005). A high organic loading rate and short retention time during the wet season are the main problems with respect to leachate treatment (Gülşen and Turam, 2005). Tränkler et al., (2005) reported that if leachate treatment plants were designed to handle the average leachate quality only, plants would occasionally be overloaded due to high discharge of leachate during certain periods of the year. Effective and successful treatment strategies require that the seasonal chemical variation in leachate composition is monitored. Among the various leachate treatment methods, reliance on natural attenuation is one of the possible strategies for pollutants in groundwater and leachate plumes (Lorah and Olsen, 1999; Christensen et al., 2001; Baun et al., 2003). Natural attenuation refers to naturally occurring physical (e.g. dilution), physico-chemical (e.g. sorption, ion exchange), chemical (e.g. precipitation) and biological (e.g. degradation) processes that can reduce concentrations of pollutants downgradient (Christensen et al., 2001). This method has rapidly become an acceptable strategy for the risk-based management of some types of groundwater contamination plumes (Wilson et al, 2004). However, effectiveness of natural attenuation depends on site and leachate characteristics, which require investigation before this approach can be implemented. Open dump systems are still widely used in Indonesia. The Jatibarang landfill receives 15 - 20% of waste from the city of Semarang, Central Java (1.4 million people in 2004). The city’s waste production has strongly increased over recent years. In 2002, total waste produced was 2500 m3 per day and this had increased to 4000 m3 per day by the end of 2005 (Cleansing Department of Semarang, 2006). The Jatibarang landfill has been in operation since 1992 as an open dumping system. A part of its leachate runs off via surface flow, during which leachate passes through a system of natural ponds and concrete leachate collection ponds, before entering the river Kreo (Fig. 1). Eleven kilometers downstream from the meeting point of leachate and the river, river water is collected and used as a major source of drinking water for the city of Semarang. Therefore, there is a need to understand the

Seasonal dynamics in leachate hydrochemistry and natural attenuation

37

Fig.1. Map of the Jatibarang landfill research site near Semarang, Central Java, Indonesia (A), showing the landfill and the leachate flowpath from landfill to the river Kreo (B) and sampling points along the leachate flowpath and the river (C). The white arrow in B indicates leachate flow from the landfill to S1. quality and quantify of landfill leachate, and the processes acting on it, to prevent surface water pollution and possible consequences for drinking water quality. Still little is known about landfill leachate characteristics in the tropics compared to temperate countries (Christensen et al., 2001; Kjeldsen and Christophersen, 2001; Kjeldsen et al., 2002; Baun et al., 2003), especially with regard to seasonal dynamics in landfill leachate characteristics (Chu et al., 1994) and potential natural attenuation of leachate. The differences in precipitation and moisture content between temperate zone landfill and tropical landfill

A

B

C

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may affect leachate composition. Given the highly seasonal climate conditions in Central Java, the fluxes of organic compounds in the leachate are expected to fluctuate seasonally, and so will their long-term impact on the quality of river water. In order to obtain insight into the strength of natural attenuation processes occurring in surface run-off between landfill and the river, leachate composition was monitored over time at selected sampling locations at the Jatibarang landfill (Fig. 1). The aims of this research were: (a) to identify seasonal landfill leachate characteristics, (b) to monitor changes in concentrations of the major leachate pollutants along the collecting ponds and at places up and downstream of the river, and (c) establish the processes contributing to changes in leachate quantity and quality, with emphasis on biodegradation. 2. Material and Methods 2.1. Field site description The Jatibarang landfill (Semarang, Central Java, Indonesia, coordinates 7o01’24.95” S, 110o21’32.23” E) has a surface area of 44.5 ha and is located in a hilly area at an altitude between 62 and 200 m above see level, with slopes up to 25 %, and 13 km from the city of Semarang. The landfill has been in operation as an open dump since 1992. It receives about 650 to 700 ton waste per day from the city of Semarang. A local community of villagers live close to the landfill and make a living from scavenging and recycling valuable materials, as well as keeping cattle that feed on the waste. Landfill leachate flows from S0 to S7 before entering the river Kreo (Fig. 1). The first leachate pond (at S1) is a natural pond. Between S2 and S7 the leachate flows through a man-made collection system, made of concrete. The leachate is collected in six connected, concrete ponds (at S5, S6) before entering the river. Annual precipitation is around of 2000 mm, air temperature ranges from 25 oC to 36 oC, air humidity ranges from 62% to 84%, soil permeability was 3.32 x 10-5 cm per second (Semarang Meteorological and Geophysical Agency, 2005).

2.2. Leachate sampling Samples were taken along the surface flow path of leachate from the landfill to the river Kreo (S1 to S7) and at four locations in the river Kreo (S8 to S11; Fig. 1). Electrical conductivity was measured in sections of the river and sampling positions downstream of the leachate inflow (S8, S10, S11) were in the part of the river with the highest electrical conductivity. Samples from S1 were used to identify the primary leachate characteristics because S1 is closest to the landfill. Leachate characteristics from other locations were used to obtain field evidence for attenuation of pollution. Sampling was conducted monthly from December 2003 to December 2004, and continued in 2005, but then only at the start and end of the wet and dry seasons. Leachate flow rates were determined for the leachate stream entering at S1. The leachate flow was measure manually using a PVC-pipe (2 m long, diameter 20 cm), the time taken to produce a fixed volume of leachate (2 L) was measured with a stopwatch. This was done 2 – 3 times per month. Rainfall data were provided by the Semarang Meteorological and Geophysical Agency. Local meteorological field data on climate could only be collected for the first four months after December 2003, thereafter it was discontinued due to interference by local villagers. For

Seasonal dynamics in leachate hydrochemistry and natural attenuation

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the four month period of measurements a strong correlation with the data reported by the nearby located Semarang Meteorological and Geophysical Agency (SMGA) was observed (data not shown). The SMGA data were used as representing the Jatibarang climate throughout.

2.3. Laboratory tests on leachate attenuation Laboratory batch tests of leachate attenuation were conducted in triplicate. Plastic containers (22 L) were filled with 20 L of leachate obtained from sampling location S1 (Fig. 1). Containers were tightly closed to avoid contact with air and rain water, and were incubated outdoors for two years. 50 ml samples were taken monthly to measure pH, electrical conductivity, alkalinity, Chemical Oxygen Demand (COD), and concentrations of calcium, ammonium, nitrate, nitrite, iron (II) and sulfate. In order to avoid infiltration of oxygen, the containers were flushed with nitrogen gas during sampling. 2.4. Analytical procedures Temperature, pH, dissolved oxygen and electrical conductivity were determined directly on site. The pH of leachate was measured using a pH meter (EcoScan, Singapore). Temperature and dissolved oxygen were measured using a HANNA instrument HI9143 DO meter. Electrical conductivity was measured using a TRANS instrument conductivity meter. Samples for determination of other leachate parameters were collected in 1 L clean polyethylene bottles and transported to the laboratory. COD, Biological Oxygen Demand (BOD), ammonium, nitrate, nitrite and phenol concentrations were measured within 6 h of sampling, while samples for sulfate, iron(II) calcium, alkalinity and organic nitrogen were stored at 4oC (no longer than 48 hours) before analysis. Samples for COD, phenol, calcium, ammonium, iron(II) and heavy metal analysis were filtered over 0.45 μm filters and conserved in 50 ml-PE bottles containing 0.4 ml concentrated nitric acid. Samples for organic nitrogen was filtered over 0.45 μm filters and conserved in 50 ml-PE bottles containing 0.5 ml HCl (18%) (Van Breukelen et al., 2003). Samples for BOD, chloride, sulfate, nitrite, nitrate and alkalinity were not filtered and kept in 50-ml PE bottles. These parameters were analyzed according to standard methods (APHA, 1998). Heavy metals were measured using spectrophotometric methods (Hach Lange GMBH, Düsseldorf, Germany). 2.5. Total coliform bacteria test A total coliform test was conducted for S8, S9, S10 and S11. Sterile PE bottles were filled with 50 ml samples from each location. Bottles were transferred on ice to the laboratory, where samples were directly tested for the presence of coliform bacteria according to standard methods (APHA, 1998). 2.6. Statistical analysis The data were analyzed using t-tests and one-way analysis of variance (ANOVA).Tukey’s multiple comparisons test was used to compare means of leachate quality among seasons and different sampling sites. Statistical computations were undertaken with SPSS for Windows version 13.

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Fig.2. Monthly precipitation (bars) and leachate flow rates (dots) at the Jatibarang landfill, Semarang, Indonesia 2.7 Calculations The dilution factor of leachate was calculated based on chloride concentrations, according to Baun et al. (2003): FM = (ClL – ClB) / (ClM – ClB). where F is the dilution factor at a particular sampling location, , ClM is the chloride concentration at this location, ClL is the chloride concentration corresponding to the upstream reference location and ClB is the background chloride concentration for non-polluted water which mixes with the leachate. In order to calculate dilution along the flow path from S1 to S7 (Figure 1), S1 was taken as the reference location (ClL) and springwater as the background chloride concentration (ClB, 22 mg L-1). To calculate the dilution by the river, the chloride concentration at S9 (upstream from the location where leachate mixes with the river: Figure 1) was taken as background ClB. Chloride concentrations were also used to predict concentrations of other chemicals along the flow path to the river, when dilution would be the sole attenuating factor: XM,expected, = XL / (ClL /ClM), where XM,expected is the expected concentration at sampling location M for chemical X, and XL is its concentration at S1. The hydraulic retention times (HRT) for leachate ponds were calculated based on the volume of the ponds (at S1, S5, S6; see figure 1) divided by the rate of leachate flow entering the ponds.

0

100

200

300

400

500

600

700

800

Dec-03

Apr-04

Aug-04

Oct-04

Dec-04

Apr-05

Aug-05

Oct-05

Dec-05

Rai

nfal

l (m

m m

onth1 )

0

20

40

60

80

100

120

140

160

180

200

Leac

hate

flow

(m3 d

-1)

wet dry wet dry wet

Seasonal dynamics in leachate hydrochemistry and natural attenuation

41

Table 1. Jatibarang leachate characteristics during the dry and wet season, at sampling location S1 (see Figure 1). Season Wet (n =11) Dry (n= 7) Significant Parameter Unit Average (range) Average (range) differences in

mean* General

pH 8.2 (7.5- 8.6) 8.3 (8.1- 8.5) NS Temperature oC 31.6 (30.7- 34.0) 30.4 (27.5- 31.7) NS

Electrical conductivity

mS cm-

1 11.7 (9.5-12.4) 12.4 (12.2- 13.0)↑ ** (p = 0.01)

Alkalinity mg L-1 3570 (2546-4406) 3466 (1227- 4453) NS Chloride mg L-1 2211 (1090-2732)↑ 1551 (1103-2249) ** (p = 0.02)

Calcium mg L-1 202 (139-393) 460 (216- 908)↑ ** (p = 0.029) Redox-related components Dissolved Oxygen mg L-1 0.04 (0.0-0.08) 0.2 (0.05 – 0.49)↑ ** (p = 0.015) Nitrate mg L-1 nd nd Nitrite mg L-1 nd nd Iron (II) mg L-1 12.2 (0.39-24.5) 15.7 (6.6-23.7) NS Sulfate mg L-1 29.9 (19.7-40.4) 46.6 (41.1-52.4)↑ ** (p = 0.04) Organic matter BOD mg L-1 303 (114-1100) 435 (213-1570) ↑ *** (p = 0.004) COD mg L-1 1260 (1062-2856) 1883 (1356-3728)↑ *** (p = 0.004) Phenol mg L-1 1.0 ( 0.15-2.5) 1.6 (0.9-2.5) NS N-organic mg L-1 255 (40-369) 421 (174-543)↑ ** (p = 0.027) N-Ammonium mg L-1 580 (376-772) 678 (408-929)↑ ** (p = 0.037)

*Student t-test. NS: not significant; ** : p < 0.05; *** : p < 0.01; ↑ indicates the season in which the particular value was significantly higher than the other season

3. Results 3.1. General leachate characteristics Flow rates through the surface leachate collecting system at Jatibarang (Fig. 1) were clearly influenced by seasonal precipitation (Fig. 2); a significant correlation between rainfall and leachate flow rate was observed (r = 0.949, p < 0.01). The seasonal variation in rainfall influenced the residence time of leachate in the natural pond at S1; in the dry season the hydraulic retention time was on average 63 days, while during the wet season it was 16 days at the lowest. Primary leachate characteristics were determined for location S1 because this location is close to the landfill (Fig. 1). Table 1 shows the average, minimum and maximum values for a number of leachate characteristics during the investigated dry and wet seasons. Overall,

Chapter 2

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ondu

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wet dry wet dry wet

A

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Seasonal dynamics in leachate hydrochemistry and natural attenuation

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Fig. 3. Seasonal variation in the average values (with standard deviation) of EC (A), BOD (B), COD (C) and NH4-N (D) concentrations in Jatibarang landfill leachate, at sampling location S1 (see Fig. 1). leachate parameters were significantly influenced by seasonal precipitation, except for pH and alkalinity. The pH of landfill leachate ranged from 7.5 to 8.6. Electrical conductivity (EC), fluctuated over time, especially during the wet season, but became more stable during the dry season (Fig.3A). The EC values in the wet season were significantly (p<0.05) lower than in the dry season, probably due to dilution by rainfall. Alkalinity of leachate was very

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Chapter 2

44

high (maximum is 4453 mg L-1), and was not significantly influenced by season. Chloride and calcium were also found at high concentrations. Calcium concentrations were significantly higher during the dry season, while chloride was significantly lower. Microbial activity in landfill leachate depends on the availability of electron acceptors. Measurement of redox parameters showed that the landfill leachate was strongly anaerobic at S1 (Table 1). Dissolved oxygen in leachate was very low (< 1 mg L-1) during both wet and dry seasons. Nitrate and nitrite were never detected during two years of sampling. However, iron(II), the product of iron(III) reduction, and sulfate were found. Iron(II) concentrations did not significantly differ between wet and dry seasons, while sulfate was significantly higher during the dry season. Oxidisable organic matter-related components (measured as chemical and biological oxygen demand; COD and BOD) and N-organics were present in significant concentrations in the landfill leachate (Table 1). In general, the average concentrations of these components were up to 200 mg L-1 and changed with the seasons, except for phenolics. BOD and COD concentrations were higher during the dry season than during the wet season (Fig. 3B, C). Also the concentrations of ammonia, a product of organic matter degradation, were significantly (p<0.05) higher during the dry season (Fig 3D).

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Seasonal dynamics in leachate hydrochemistry and natural attenuation

45

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Chapter 2

46

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Seasonal dynamics in leachate hydrochemistry and natural attenuation

47

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Chapter 2

48

Figure 4. Seasonal changes in the average (with standard deviation) of chloride (A), EC (B), calcium (C), BOD, (D), COD (E), ammonium (F), iron (G), and sulphate (H) concentrations along the flow path from the Jatibarang landfill to river Kreo. Measured concentrations are shown by bars (white, wet season; dark, dry season), while predicted values (assuming dilution as single factor in natural attenuation) are indicted by symbols (white circles, wet season; black circles, dry season), connected by lines. 3.2. Natural attenuation of landfill leachate at Jatibarang Natural attenuation comprises all naturally occurring physical, chemical and biological processes that reduce the toxicity of contaminants (Lorah and Olsen, 1999). The time available for natural attenuation to occur at Jatibarang was short: the hydraulic retention time (HRT) for the natural pond (S1) during the dry season was clearly higher (45 – 63 days) than during the wet season (16 – 34 days). HRTs in the concrete collection ponds (S6) during wet and dry seasons were only 2 – 4 days and 1.5 – 3 days respectively. Nevertheless, concentrations of organic matter (BOD, COD), nitrogen (ammonia) and redox-related compounds (iron(II), sulphate) showed changes along the flowpath from landfill to river; they strongly decreased with increasing distance from the landfill (Fig. 4). The average ammonium concentration decreased between S1 and S7 by 43% and 55% during the wet and the dry seasons (Fig. 4F), respectively. In the river the concentrations were lowered by 99% and 98% during the wet and the dry season, respectively. Similar observations were also made for other leachate parameters. The average COD concentrations decreased between S1 and S7 to 31% and 43% during the wet and dry seasons, respectively, and BOD with 48% and 61%, respectively (Fig. 4D, E). EC decreased in the ponds and in the river by up to 90% (Fig. 4B, Table 2). Concentrations of leachate components in general dropped strongest at sampling points S3 and S4 (Fig. 4). Spring water was led into the leachate collection system between these sampling points during the dry season in order to maintain sufficient volume of leachate in the collection ponds (Fig. 1). This practise was conducted in order to avoid that the concrete

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Seasonal dynamics in leachate hydrochemistry and natural attenuation

49

collections ponds (S5, S6) would dry out and start to break and leak. During the wet season the area near S3 and S4 occasionally flooded, and rainwater also entered the leachate collecting system. The importance of dilution of leachate by springwater and rainfall in the decreases in leachate components along the flow path were derived from calculations by considering chloride as a conservative tracer (section 2.7). The dilution factor in the ponds in the wet and the dry season ranged from 2 to 5, respectively. By correcting for dilution, via normalisation based on the conservative tracer chloride (section 2.7), we calculated at each sampling point the expected concentration of the leachate components (lines in Fig. 4). Comparison to the observed data (bars in Fig. 4) revealed that dilution was the major attenuating factor. Calculated values corresponded well to the measured concentrations between S1 and S4. In the leachate collection ponds (S5, S6) the calculated concentrations were even lower than the measured concentrations. This suggests net production, or release, of leachate components during leachate flow from the landfill to the river. The dilution factor of leachate by river water at S8 was around 140. In general, the decrease in concentrations between landfill and river (S1 – S7) was much smaller than the dilution upon entry of leachate into the river (Fig. 4, Table 2). Table 2. River quality at the meeting point of leachate and the river (S8) and along the river up (S9) and down stream (S10 – S11), in comparison to leachate composition (S7). For sampling locations see Fig. 1.

Parameter

In leachate effluent

In the river

S7 S8 S9 S10 S11 pH 8.4 ± 0.3 a 8.4 ± 0.3 a 8 ± 0.6 a 8 ± 0.4 a 8 ± 0.5 a EC (mS cm-1) 4.5 ± 1.5 a 3.2 ± 1.3 a 0.14 ±0.03 b 0.17 ±0.1 b 0.14±0.04

b DO (mg L-1) 4 ± 0.9 a 6 ±1.01 b 7.2 ± 1 b 7 ± 0.7 b 6.1 ± 1.6 b BOD (mg L-1) 264 ±69 a 56±4.6 b 13±1.3 b 17±2.4 b 19±10 b COD (mg L-1) 961± 35.1 a 343±45 b 64±6.4 c 70±5.2 c 54±40 c N-NH4 (mg L-1) 375 ± 134 a 90 ± 53 b 3 ± 2.1 c 2.5 ± 1.3 c 2.2 ± 0.5 c N-NO3(mg L-1) 1.7 ± 3.7 a 0.7 ± 0.5 a 0.9 ± 0.8 a 1 ± 0.7 a 1 ± 1.01 a N-NO2(mg L-1) nd 0.32±0.27 a 0.05±0.02 a 0.1±0.07 a 0.11±0.09

a Cl (mg L-1) 1131 ± 297 a 396±156 b 24± 20 c 32± 30 c 17±6 c Ca (mg L-1) 161 ± 38 a 80 ± 23 b 21± 3 c 20 ± 5 c 18± 3 c Fe2+ (mg L-1) 4.2 ± 3.3 a 2 ± 1.6 b 0.8 ± 1.8 c 0.8 ± 1.7 c 0.9 ± 2 c SO4 (mg L-1) 19 ± 7 a 7 ± 6 b 0.7±0.4 c 0.8±0.2 c 0.7±0.3 c All data are presented as means (n=3). Significant differences (p < 0.05) between sites are indicated by a different character behind the standard deviation values. nd, not detected.

Chapter 2

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Minor changes in redox conditions were observed along the flow path. Although nitrite and nitrate concentrations were never above the detection limit (0.01 mg-N L-1) at S1, nitrite and nitrate were occasionally detected at S7. Nitrite was found once at a concentration of 0.25 ± 0.01 N-mg L-1, while nitrate was detected four times, with a maximum concentration of 8.4 ± 0.03 N-mg L-1. This is probably caused by nitrification because dissolved oxygen was also occasionally detected in S7 (oxygen was only detected five times during the sampling period and the maximum concentration was 4.5 mg L-1), suggesting that oxygen was introduced by mixing processes and diffusion while the leachate flowed from S1 to S7. 3.3. Impact of landfill leachate on river water quality To assess the impact of leachate on river water quality, selected sampling points were monitored for one year (December 2003 to December 2004). Samples were taken at S7, which is along the leachate flowpath and closest to the river, the meeting point between leachate and river (S8) and sampling points in the river that were up (S9) and down-stream (S10, S11) of the leachate entry point. Significant differences in dissolved oxygen, BOD, COD, ammonia, chloride, calcium, iron(II) and sulphate but not in pH, EC, nitrite and nitrate were found between S7 and S8 (Table 2). However, for none of the twelve properties a significant difference between river water upstream and downstream of the point where leachate entered the river was observed (Table 2). In contrast, an impact of heavy metals in the leachate on river water quality was clearly observed. Dissolved heavy metals were measured at location S7 and river water in June 2004. The concentrations were 110 µg Cd L-

1, 152 µg Cr L-1, 1.51 µg Cu L-1, 200 µg Pb L-1, 6020 µg Mn L-1 450 µg Ni L-1 and 3080 µg Zn L-1, exceeding concentrations in the river water with which it mixed (4.1 µg Cd L-1, 38 µg Cu L-1, and 140 µg Zn L-1). Average total concentrations in sediment of river Kreo downstream of the landfill were 8.25 μg Cd, 1434 μg Cu and 1162 μg Zn per g dry weight), in general much higher than upstream (8.97, 124 and 800 μg per g dry weight, respectively). Total coliform bacteria counts at the inlet of leachate into the river (S8), and several sampling points along the river (S9, S10 and S11) were (per ml); 16, 0.55, 0.54 and 0.32, respectively. 3.4. Natural attenuation potential during laboratory incubations The potential for natural attenuation of leachate components was determined in the laboratory with leachate collected from location S1 in August 2004. Changes in leachate characteristics occurred during a 2 year incubation (Fig. 5). pH slightly decreased in the first four months and increased later on up to 11.4 at the end of the experiment (Fig. 5A). The increase in pH at the end of test appears correlated with the increase in alkalinity (Fig.5A). This probably indicates a decrease in metals due to precipitation of metal sulphides, hydroxides and carbonate (Ward et al., 2005). A similar trend was observed for calcium, iron(II), and sulphate decreased during the experiment. COD concentrations decreased in the first four months but then stabilized. Ammonium concentrations increased slightly in the first two months, probably caused by protein decomposition. Nitrate and nitrite appeared when the ammonium concentration started to decline from around 500 mg L-1 to less than 40 mg L-1. Nitrite was produced at concentrations up to 90 mg-N L-1, while nitrate was present at less than 15 mg-N L-1. Oxygen concentrations were measured at the start (0.42 mg L-1) and at the end of the experiment (0.15 mg L-1), indicating anaerobic conditions. The decrease in iron(II) concentration during the first four months of incubation was probably due to precipitation as

Seasonal dynamics in leachate hydrochemistry and natural attenuation

51

siderite (FeCO3) and iron-sulphide, since precipitates were observed at the bottom of the containers. Generally, these results indicate that chemical precipitation and biological processes were occurring under anaerobic to semi-anaerobic conditions in the landfill leachate, as can be seen from the redox changes (iron(II), sulphate, nitrite and nitrate).

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Chapter 2

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Fig.5. Changes in EC (●), COD (○) (A), iron(II) (□), sulphate (■) (B), ammonium (Δ), nitrite (○) and nitrate (●) (C), pH (♦), alkalinity (○) and calcium (Δ) (D) during a two-year laboratory test.

4. Discussion 4.1. Leachate composition. Landfill leachate composition varies between landfills and within landfills (Christensen et al., 2001). Leachate composition relates to the stage of waste decomposition. Waste decomposition occurs in three major stages: an aerobic stage, an anaerobic acidogenic stage

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Seasonal dynamics in leachate hydrochemistry and natural attenuation

53

and finally, an anaerobic methanogenic stage (Kjeldsen et al., 2002). In the aerobic stage, the production of leachate is low, but leachate is heavily polluted and has a slightly neutral pH (7.0 – 7.8). In the anaerobic acidogenic stage, the leachate has a high organic volatile fatty acid content and low pH. BOD and COD concentrations are highest in this phase (Kjeldsen et al., 2002). In the methanogenic stage, the pH increases to neutral or above and the BOD/COD ratio drops to less than 0.1 (Chen, 1996; Christensen et al., 2001; Kjeldsen et al., 2002). The Jaribarang landfill leachate is classified as a typical old leachate (after 10 years of operation), with a high pH (7 – 8; Table 3). Increasing pH values indicate a shift from the acetogenic phase to the methanogenic phase (Ehrig, 1983). A pH above 7 is indicative of low acetogenic activity (Fadel-el et al., 2002). The Jatibarang leachate composition was compared to that of other tropical landfills and landfills in temperate climate zones (Table 3). The Jatibarang landfill leachate was highest in alkalinity, chloride and calcium. Redox-related components such as iron(II) and sulfate in leachate from this landfill were intermediate to those observed at other landfills, their concentrations were higher than those from Hong Kong and Taiwan but lower than in landfill leachate from Nigeria. Concentrations of ammonium and organic compounds (COD) in Jatibarang were comparable to the landfill leachate from Nigeria and higher in the leachate of the other landfills. High ammonium concentrations are commonly observed in landfill leachate; this is mainly due to decomposition of proteins in organic waste (Kjeldsen et al., 2002). Organic waste comprises about 62% of the waste dumped at the Jatibarang landfill (Cleansing department city of Semarang, 2003). Nitrite and nitrate were in general below the detection limit. This is probably due to a lack of a significant nitrifying microbial community in the leachate, which relates to the absence of oxygen, as one would expect for a landfill in the methanogenic phase.

Fig. 6. Changes in BOD/COD ratio at sampling location S1.

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54

Table 3 Leachate composition of landfills located in tropical and temperate climate zones

Jatibarang a * Hongkong b* Nigeria c * Taiwan d * Denmark e ** Netherlandsf** Canadag**

Parameter Unit Range Range Range Range average average

pH 7.5 – 8.6 7.2 – 8.4 8.03 – 8.28 7.0 – 8.5 7 6.6

EC mS cm-1 9.5 – 13.0 2.5 – 11.8 4.8 - 5.6 3.6-14.1 3

Alkalinity mg L-1 1226 – 4453 1421 – 2208

Chloride mg L-1 1090 – 2732 140 - 1103 1271 - 1606 360 260

Calcium mg L-1 138.9 – 908.4 47.2-137.5

Dissolved Oxygen mg L-1 0.0 – 0.49 1.9 – 2.1

N-NO2 mg-N L-1 nd 21.6 –179

N-NO3 mg-N L-1 nd 0.47 – 0.58

Iron(II) mg L-1 0.39 – 24.5 1.01 – 3.45 122.4 – 180.4 0.26-5.4 45+

Sulfate mg L-1 19.7 – 52.4 65.33 – 111.2 150 < 3

BOD mg L-1 113 – 1570 675.6 – 990.6 12 - 97 44 930

COD mg L-1 1061 – 3727 147 - 1590 2802 – 3066 320-1340 320 1870

Phenol mg L-1 0.15 – 2.5 23

N-organic mg-N L-1 40 – 542 72 - 177 40 - 150

N-NH4 mg-N L-1 375 – 929 65 - 883 622 – 1316 110 360 10 a. Present study; b.Chu et al., 1994; c.Aluko et al., 2003; d.Fan et al., 2005; e Kjeldsen and Christopherse, 1992; f Van Breukelen, 2003; g Henry et al., 1987. nd: not detected; * Tropical zone; ** Temperate zone

55

COD and BOD concentrations in the leachate were in the range observed for other old landfills (Kjeldsen et al., 2002). Fig. 6 show that a decrease in BOD/COD ratio occurred in the first year of sampling, but this ratio increased in the second year. There was no significant difference in ratio between the wet and dry season (0.1 – 0.5 and 0.1 – 0.3 respectively). A low BOD/COD ratio indicates stability of the leachate (Fan et al., 2005; Kjeldsen et al., 2002). The increase in BOD/COD ratio in the second year of sampling is possibly due to the mixing of fresh landfill material with old leachate. Jatibarang is still actively used and due to space and finance limitations there is no proper separation between old and new waste. Generally, the BOD/COD value in the Jatibarang landfill leachate indicates a moderately stable landfill, in between the end acidogenic and early methanogenic stage (Fadel-el et al., 2002).. Iron(II) was found during wet and dry seasons in relatively high concentrations, while oxygen and nitrate were absent, suggesting that iron reduction occurs (Christensen et al., 2001, Van Breukelen et al., 2003). 4.2. Seasonal effects Not unexpectedly, the highest leachate flow rates were observed during the wet season, resulting in a high flux of pollutants. The average loads of ammonium and COD during the wet season were 79 kg-N per day and 222 kg-COD per day, respectively. These values were on average three times higher than in the dry season (24 kg-N per day and 68 kg-COD per day, respectively). In the wet season, rain water that percolates through the waste dump will extract adsorbed and soluble compounds, producing a large volume of diluted leachate (Chu et al., 1994; Chen, 1996). During the dry season in general a smaller volume of more concentrated leachate is produced because there is only minor input of rain water and more evaporation (Ehrig, 1983; Lema et al., 1988; Chu et. al., 1994; Fungaroli and Steiner, 1971). However, in some cases no significant correlation between season and leachate quality has been observed or even higher concentrations of leachate components during the wet season (Äkenson and Nilsson, 1997; Jaspers et al., 1985). Also, Fan et al. (2005) observed that organic matter (COD, BOD), chloride and calcium concentrations were higher in leachate from a landfill in Taiwan during the wet season than during the dry season. The primary Jatibarang leachate at location S1 became more concentrated during the dry season due to a low input of rain water and more evaporation, as judged from its darker colour and the decreasing leachate volume in the stream. In general, most chemical characteristics such as EC, ammonium and organic compounds (COD) were higher during the dry season with the exception of chloride and BOD. There were no significant differences in pH, alkalinity and iron(II) over the seasons. This indicates that the seasonal effect on leachate quality at Jatibarang is complex. Changes in chemical concentrations of leachate did not only occur by simple dilution due to seasonal precipitation and evaporation, but also other factors such as waste composition, landfill management practice and internal landfill processes are probably influential (Chen, 1996; Fadel-el et al., 2002). 4.3. Field evidence for natural attenuation Leachate contaminants decreased in concentration along the flow path from landfill to river which provides field evidence for the occurrence of natural attenuation. The critical point in natural attenuation is to distinguish between dilution and biodegradation (Baun et al., 2003). Dilution appears to be the predominant process at our site, the decreases in concentrations of non-degradable components such as chloride along the flow path are indicative of dilution

Chapter 2

56

(Christensen et al., 2001). During the wet season rainwater entered the ponds while during the dry season spring water was directed to the ponds at S6 in order to maintain a constant leachate volume. The dilution factor during the dry season was higher than during the wet season. Many organic pollutants in landfill leachate are well-degradable under anaerobic conditions (Christensen et al., 2001). However, although our laboratory experiment clearly revealed strong decreases in concentrations of organic leachate contaminants under limited oxygen conditions, degradation rates were low, certainly in light of the residence time of the leachate collecting system at Jatibarang (<63 days at S1, <4 days at S5). In the laboratory experiment, the COD concentration decreased by less than 25% over a period of 70 days. Over a period of 4 months it decreased by 65%, after which it remained stable. Humic substances (humic and fulvic acid) are the main organic components in leachate (Calace and Petronio, 1997). They are recalcitrant to degradation and can sorb heavy metals, thus a high concentration of non-degradable humic substances might compromise the treatment of landfill leachate. The laboratory experiments also revealed the removal of ammonium, which might be due to nitrification, nitrate reduction and/or anaerobic ammonium oxidation (anammox). Most biological nitrogen removal processes are carried out by the combination of aerobic nitrification, anoxic denitrification and anaerobic ammonium oxidation processes (Chen, 1996; Fux et al., 2002; Pelkonen et al., 1999; Jokella et al., 2002). Under aerobic conditions ammonium can be sequentially oxidized to nitrite and nitrate by nitrifiers. When oxygen is depleted, nitrate can be converted to nitrite and finally to dinitrogen gas by denitrification. Also, when nitrite is present under anaerobic conditions, ammonium can be oxidized with nitrite as electron acceptor to dinitrogen gas (anammox) (Mora et al., 2004). Although in our experiment 73% of ammonium is removed within the first 4 months, during the time equalling the retention time of the leachate collecting system (~70 days) no significant ammonium degradation was observed (Fig. 5C). The occurrence of nitrate in the laboratory experiment is probably due to nitrifier activities, suggesting the entrance of some oxygen into the system despite the fact we did not measure significant oxygen concentrations. Thus, reliance on intrinsic biodegradation in the leachate collecting ponds does not pose an acceptable treatment at Jatibarang. The biodegradation processes require considerable time, while the average retention time in the leachate collection ponds was only 48 hours, even in the dry season. In fact, a net production of leachate components appears to occur. Predicted concentrations of leachate components were lower than the measured concentrations. This may relate to the observed, occasional growth of algae in the leachate ponds (S5, S6), which via photosynthesis could introduce organic matter into the leachate, and to exchange reactions between leachate and the concrete over which it flowed. 4.4. Long-term impact on river Concentrations of leachate components in the river were below standards set by the Indonesian Ministry of Health (2003), due to strong dilution. However, for the long term, accumulation of leachate pollutants in the river ecosystem must be taken into account. Heavy metals already appear problematic. Heavy metals were present in Jatibarang leachate, in high concentrations between 0.11 mg L-1 (Cd) to 6.02 mg L-1 (Mn). Their concentrations were much higher than those observed in the river, which already exceeded Indonesian regulations (Indonesian Ministry of Health, 2003). A high content of heavy metals in river sediments was

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also observed. Landfill leachate contains high numbers of bacteria, in general higher than found in pristine water (Christensen et al., 2001). Pathogenic bacteria such as faecal coliforms and streptococci are found in leachate (Sadek and Fadel-El, 2000). This study showed that the number coliform bacteria in the river were lower than national regulatory standards in Indonesia for drinking water (< 2.2 counts per ml) (Indonesian Ministry of Health, 2003), but not for the leachate (S8). Regarding the presence of heavy metals in the sediment and the coliform bacteria in the river, one has to consider the long-term impact of heavy metal toxicity on aquatic organisms and the impact of pathogenic bacteria on human health risk which using the water for drinking water supply. A significant negative correlation between abundance of benthic macroinvertebrates and metal concentrations in the sediment of the river Kreo has been observed (data not shown).To prevent the risk for the Semarang population and to safeguard the river Kreo as a source of drinking water, it is imperative to prevent pollution of the surface water by the landfill. For this reason, and due to the limited intrinsic biodegradation during surface leachate flow from landfill to the river, the Jatibarang landfill management system should be improved to diminish landfill leachate quantity and quality. The residence time of the leachate in the leachate collecting system could be increased to allow for more biodegradation and precipitation, by enhancing the volume capacity of the leachate collecting system. 5. Conclusions Generally, the concentration of Jatibarang landfill leachate components in primary leachate were higher than observed at other tropical landfills. The Jatibarang landfill leachate quantity and quality depends strongly on season and highest during the wet season. Significant natural attenuation occurred by dilution; biodegradation did hardly take place along the flow path of leachate to the river even though laboratory experiments indicate the potential for substantial removal of ammonium and BOD. No significant impact of landfill leachate on river water quality was observed, however the presence of heavy metals and coliform bacteria need to be considered on the long-term. Although dilution can decrease the concentration of pollutants in landfill leachate, it is not an optimal attenuation mechanism because it does not decrease the absolute load of pollutants into the river. The Jatibarang landfill management system should be improved to diminish landfill leachate quantity and quality. This can be achieved by increasing the residence time of leachate in the leachate collecting system by increasing its volume capacity. Acknowledgements The authors would like to thank the V. Rutgers fund of the VU University Amsterdam, The Netherlands, for funding the work described here. Maartje Kunen and F.S Papilaya are acknowledged for drawing Figure 1. Special thanks go to Alfons Maramis for providing data on heavy metals. We also thank Oky Wisnu, Christine Setyaninggrum and Ocon Totoda for their assistance in the laboratory and field work.

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Åkesson, M., and Nilsson, P., 1997. Seasonal changes of leachate production and quality from test cells. Journal of Environmental Engineering 123, 892-900.

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Calace, N., and Petronio, B.M., 1997. Characterization of high molecular weight organic compounds in landfill leachate: humic substances. Journal of Environmental Science Health 32, 2222-2224.

Chen, H.P., 1996. Assessment of leachates from sanitary landfills: Impact of age, rainfall and treatment. Environmental International 22, 225-237.

Chian, E.S.K., and Dewalle, F.B, 1976. Sanitary landfill leachate and their treatment. Journal of The Environmental Engineering Division 120, 411-431.

Christensen, T.H., Kjeldsen, P., Bjerg, P.L., Jensen, D.L., Christensen, J.B., Baun, A., Albrechtsen, H.-J., Heron, G. 2001. Biogeochemistry of landfill leachate plumes. Applied Geochemistry 16, 659-718.

Chu, M.L., K.C. Cheung., M.H. Wong, 1994. Variation in the chemical properties of landfill leachate. Environmental Management 18, 105-117.

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Ehrig, H.J., 1983. Quality and quantity of sanitary landfill leachate. Waste Management and Research 1, 53-68.

Fadel-el, M., E. Bou-Zeod., Chahine, W., Ayali, B., 2002. Temporal variation of leachate quality for pre-sorted and baled municipal solid waste high organic and moisture content. Waste Management 22, 269-282.

Fan, J.H., Shu Y.H., Yang, S.H., Chen, C.W., 2005. Characteristic of landfill leachate in central Taiwan. Science of the Total Environment 361, 25-37.

Fungaroli, A.A., and Steiner, R.L., 1971. Laboratory study of the behaviour of a sanitary landfill. Journal of Water Pollution Control Fed 42, 252-267.

Fux, C., Boehler, M., Huber, P., Brunner, II, Siegrist, H., 2002. Biological treatment of ammonium-rich wastewater by partial nitritation and subsequent anaerobic ammonium oxidation (ANAMMOX) in pilot plant. Journal of Biotechnology 99, 295-306.

Gülşen, H., and Turam, M., 2005. Anaerobic treatability of sanitary landfill leachate in a fluidized bed reactor. Turkish Journal of Engineering. Environmental Science 28, 297-305.

Henry, J.G., Prasad, D., and Young, H., 1987. Removal of organics from leachates by anaerobic filter. Water Research 21, 1395-1399.

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Jaspers, S.E., Atwater, J.W., and Mavinic, D.S., 1985. Leachate production and characteristics as a function of water input and landfill configuration. Water Pollution and Research Journal of Canada 20, 43-56.

Johansen, J.O., and Carlson, A.D., 1976. Characterization of sanitary landfill leachates. Water Research 10, 1129-1134.

Jokella, J.P., Kettunen, R.H., Sormunen, K.M. Rintala, J.A., 2002. Biological nitrogen removal from municipal landfill leachate: low-cost nitrification in biofilters and laboratory scale in-situ denitrification. Water Research 36, 4079-4087.

Kjeldsen, P., 1993. Groundwater pollution source characterization of and old landfill. Journal of Hydrology 142, 349-371.

Kjeldsen, P., and Christophersen, M.., 2001. Composition of leachate from old landfills in Denmark. Waste Management and Research 19, 249-256.

Kjeldsen, P., Bariaz, A.M., Rooker, P.A., Baun, A., Anna Ledin and Christensen, H.T., 2002. Present and long-term composition of MSW landfill leachate: A review. Critical review in Environmental Science and Technology 32, 297-363.

Lema, M.J., Mendez, R., Blazquez, R., 1988. Characterization of landfill leachates and alternatives for their treatment: a review. Water Air and Soil Pollution 40, 223-250.

Lorah, M.M., and Olsen, D.L., 1999. Natural attenuation of chlorinated volatile organic compounds in a fresh tidal wetland. Water Resources Research 35, 3811-3827.

Miyajima, T., Wada, E., Hanba, T.Y., Vijarnsorn, P, 1997. Anaerobic mineralization of indigenous organic matters and methanogenesis in tropical wetland soils. Geochemical et Cosmochimica Acta 61, 2739-3751.

Mora, D.A., Arrojo, B., Campos, L.J., Corral, M.A., Méndez, R., 2004. Improvements settling properties of Anammox sludge in an SBR. Journal of Chemical Technology Biotechnology 79, 1417-1420.

Nielsen, P.H.,Christensen, T.H., 1994.Variability of biological degradation of aromatic hydrocarbon in an aerobic aquifer determined by laboratory batch experiments. Journal of Contaminant Hydrology 15, 305-320.

Pelkonen, M., Kotro, M., Rintala, J., 1999. Biological nitrogen removal from landfill leachate: a pilot-scale study. Waste Management and Research 17, 493-497.

Sadek, S. and Fadel-El, M., 2000. The Normandy Landfill: A case study in solid waste management. Journal of Natural Resources Life Science Education 29, 155-161.

Semarang Meteorological and Geophysical Agency, 2005. Daily rainfall data in Central Java. Semarang, Indonesia.

Tränkler, J., Visvanathan, C., Kuruparan, P., Tubtimthai, O., 2005. Influence of tropical seasonal variation on landfill leachate characteristics-results from lysimeter studies. Waste Management 25, 1013-1020.

Van Breukelen, B.M., Röling, W.F.M., Groen, J., Grifioen, J., Van Verseveld, H.W., 2003. Biogeochemistry and isotope geochemistry of a landfill leachate plume. Journal of Contaminant. Hydrology 66, 245-268.

Van Breukelen, B.M., 2003. Natural attenuation of landfill leachate: a combined biogeochemical process analysis and microbial ecology approach. Ph.D. thesis. VU University Amsterdam, The Netherlands.

Ward, M.L., Bitton, G., Townsend, T., 2005. Heavy metal binding capacity (HMBC) of municipal solid waste landfill leachate. Chemosphere 60, 206-215.

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3 Microbial nitrogen transformation potential in surface run-off leachate from a tropical landfill

Mangimbulude, J.C., Van Straalen, N.M. Röling, W.F.M., 2012

Waste Management 32, 77-87

Abstract Ammonium is one of the major toxic compounds and a critical long-term pollutant in landfill leachate. Leachate from the Jatibarang landfill in Semarang, Indonesia, contains ammonium in concentrations ranging from 376 to 929 mg-N L-1. The objective of this study was to determine seasonal variation in the potential for organic nitrogen ammonification, aerobic nitrification, anaerobic nitrate reduction and anaerobic ammonium oxidation (anammox) at this landfilling site. Seasonal samples from leachate collection treatment ponds were used as an inoculum to feed synthetic media to determine potential rates of nitrogen transformations. Aerobic ammonium oxidation potential (< 0.06 mg N L-1 h-1) was more than a hundred times lower than the anaerobic nitrogen transformation processes and organic nitrogen ammonification, which were of the same order of magnitude. Anaerobic nitrate oxidation did not proceed beyond nitrite; isolates grown with nitrate as electron acceptor did not degrade nitrite further. Effects of season were only observed for aerobic nitrification and anammox, and were relatively minor: rates were up to 3 times higher in the dry season. To completely remove the excess ammonium from the leachate, we propose a two-stage treatment system to be implemented. Aeration in the first leachate pond would strongly contribute to aerobic ammonium oxidation to nitrate by providing the currently missing oxygen in the anaerobic leachate and allowing for the growth of ammonium oxidisers. In the second pond the remaining ammonium and produced nitrate can be converted by a combination of nitrate reduction to nitrite and anammox. Such optimization of microbial nitrogen transformations can contribute to alleviating the ammonium discharge to surface water draining the landfill. Key words: municipal solid waste; tropical landfill leachate; natural attenuation; nitrogen transformation; anammox

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1. Introduction

Landfill leachate generally contains high concentrations of inorganic nitrogen compounds such as ammonium, nitrite and nitrate (Christensen et al., 2001; Kjeldsen et al., 2002; Kim et al., 2006, Mangimbulude et al., 2009). Inorganic nitrogen compounds may leach to groundwater and surface water and pose serious risks to ecosystems and human health. Ammonium is the primary cause of acute toxicity of municipal landfill leachate (Kim et al., 2006; Clěment and Merlin, 1995; Ernst et al., 1994). Nitrogen transformations in landfills, which include ammonification of organic nitrogen, nitrification, denitrification, ammonia volatilization, and anaerobic ammonium oxidation, are mainly driven by microorganisms (Berge and Reinhart, 2005). Transformations start with ammonification of organic landfill components, in particular proteins (Jokela et al., 2002; Berge and Reinhart, 2005). The conversion of protein-nitrogen to ammonium by heterotrophic bacteria occurs in two steps, first an enzymatic hydrolysis by aerobic and anaerobic microorganisms releasing free amino acids and secondly, de-amination of the amino acids to release ammonium. Ammonium can next be oxidized to nitrate by a process called nitrification. Nitrification takes place under aerobic conditions in two steps, each performed by specialized groups of autotrophic microorganisms; ammonia oxidation leading to nitrite followed by nitrite oxidation to nitrate. Nitrification is slow and suppressed by high concentrations of ammonium and organic matter (Joo et al., 2005). Subsequently, the products of nitrification might be reduced to nitrogen gas by heterotrophic organisms (Betlach and Tiedje, 1981). This process of denitrification takes place mainly under anaerobic conditions and requires an electron donor which can be organic material or reduced inorganic compounds such as sulfide, ferrous iron, methane or hydrogen (Van Loosdrecht and Jetten, 1998; Raghoebarsing et al., 2006). Dinitrogen gas is also the final product of anaerobic ammonium oxidation (anammox) which involves the oxidation of ammonium with nitrite as electron acceptor by obligately anaerobic autotrophic bacteria of the bacterial phylum Planctomycetes (Strous et al., 1999); hydroxylamine and hydrazine have been identified as intermediates of the anammox process (Van de Graaf et al., 1995; Egli et al., 2001). Biological nitrogen removal in landfill leachate is still facing difficulties because it is suppressed, not only by ammonium but also by inhibitors such as volatile fatty acids, phenolics and humic acids (Berge and Reinhart, 2005; Dyreborg et al., 1995; Takai et al., 1997; Kulikowska and Klimiuk, 2004). Ammonium concentrations (376 to 929 mg-N L-1) in anaerobic surface run-off leachate from a tropical landfill, Jatibarang (Semarang, Indonesia), always exceeded nitrate concentrations, indicating that ammonium availability itself was not limiting nitrification (Mangimbulude et al., 2009). Nitrate and nitrite concentrations were seldom above the detection limit (0.1 mg/l). Similar observations have been made for landfills in temperate climate zones (Kjeldsen et al., 2002). Availability of oxygen is one of the most important environmental factors controlling the rate of nitrification (Prinčič et al., 1998). Goreau et al. (1980) and Kim et al. (2006) reported that dissolved oxygen concentrations below 2 mg L-1 were limiting nitrification.

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Fig. 1. Map of the Jatibarang landfill research site near Semarang, Central Java, Indonesia (A), showing the landfill and the leachate flowpath from the landfill to the river Kreo (B, C). Sampling points along the leachate flowpath and the river, sampled in a previous study (Mangimbulude et al. 2009) are indicated as S0 – S11. S0 and S6 were sampled in this study.

Nitrogen transforming processes in tropical landfill leachate have received only minor attention so far. Meteorological factors, especially rainfall, are known to vary within very large amplitudes in the highly seasonal climate of tropical countries like Indonesia (involving marked dry and wet periods). Ammonium and organic N concentrations in leachate of the Jatibarang landfill were significant influenced by rainfall (Mangimbulude et al., 2009),

A

B

C

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indicating considerable seasonal variation in microbial nitrogen transformations. The large differences in nitrogen fluxes between wet and dry seasons provide a difficulty in controlling biological nitrogen removal in landfill leachate. A better understanding of the occurrence and seasonal variation in the potential for nitrogen mineralization, nitrification, denitrification, and anammox in leachate could help to improve the management of tropical landfills, and prevent surface water quality deterioration due to uncontrolled discharge of leachate containing high concentrations of toxic ammonium. In the Jatibarang landfill, a part of the leachate runs off via surface flow, during which leachate passes through a system of natural ponds and concrete leachate collection basins, before entering a river (Fig. 1). Field measurements revealed that less than 25% of ammonium in the Jatibarang leachate collecting system was removed by biological processes (Mangimbulude et al., 2009). In the present study, we aimed to determine the potential for organic nitrogen mineralization, through ammonification, nitrification, denitrification and anammox in natural and artificial leachate collecting ponds at the Jatibarang landfill. In this way we obtained an overview of the capacity of microbial communities for inorganic nitrogen transformation which will be useful for risk assessment and the design of an appropriate treatment technology for nitrogen removal in surface run-off leachate. Samples from the field were taken at different times during the seasonal cycle and used as inocula for nitrogen-transformation experiments. Initial potential rates were determined from short-term incubations to minimize the influence of growth, while longer incubations provided information on the overall potential 2. Material and methods 2.1. Field site The Jatibarang landfill (Semarang, Central Java, Indonesia) covers 44.5 ha and is located in a hilly area at an altitude between 62 to 200 m (with coordinates 7o01’24.95” S, 110o21’32.23” E), with a slope up to 25 %, 13 km South of the city of Semarang. The landfill is in operation as an open dump site since 1992. It receives 650 to 700 tonnes of waste per day. The landfill leachate produced is collected in four connected ponds before flowing into the river Kreo (Fig. 1). The annual precipitation at the site is in the range of 2,013 to 2,215 mm, air temperature varies between 25 oC and 36 oC, air humidity is 62% to 84%, and soil permeability was measured as 3.32 x 10-5 cm s-1 (Semarang Meteorological and Geophysical Station, 2005).

2.2. Sample preparation Landfill leachate samples were taken in different seasons (August 2005 representing the dry season and January 2006 representing the wet season) and at different sites; S0 is closest to the landfill and S6 is the last of a series of leachate collecting ponds (Fig. 1). Collected leachate was transported within 2 hours to the laboratory and 100 ml aliquots of leachate were centrifuged at 10.000 rpm for 10 minutes. The supernatants were discharged and the pellets were resuspended in 100 ml salt solution (0.9% (w/v) of NaCl). To create anaerobic conditions for anaerobic transformation experiments, the salt solution containing resuspended leachate solids was flushed with nitrogen gas for 5 to 10 minutes.

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2.3. N-Arginine mineralization tests In order to determine the potential for organic-N mineralization, 2.5 ml of resuspended leachate solids was inoculated into 50 ml of medium containing 50 mg L-1 of arginine, as described by Alef and Nannipieri (1995). These culture bottles were closed with a rubber stopper, incubated under aerobic conditions (using a shaker at 100 rpm) or anaerobic conditions (flushed with nitrogen gas for 5 to 10 min, and sealed) at room temperature (28 ºC) for 3 h. Samples were taken at four different times (0 h, 0.5 h, 1.0 h and 3.0 h) and processed for analysis for ammonium. The rate of ammonium release was estimated from the slope of a linear regression applied to the ammonium concentrations plotted against time, and was expressed as mg-N per h-1 L of leachate per hour. Three replicates were conducted for each individual leachate sample. 2.4. Nitrification tests An aerobic nitrification test was applied to estimate the N-NO3 production rate (Krave et al., 2002). Sample of 2.5 ml resuspended leachate solids were added to 500 ml glass bottles containing 100 ml of sterile nitrification medium modified from Alef and Nannipieri (1995) (per liter: 6 g of CaCO3, 0.0272 g of KH2PO4, 0.0294 g of CaCl2.2H2O, 0.492 g of MgSO4.7H2O, and 0.95 g of (NH4)2SO4). The pH of the medium was adjusted to 7.5, and the inoculated medium was incubated in the dark using a rotary shaker (150 rpm) at room temperature (28 ºC) for four weeks. During incubation, 12 – 14 samples were taken over time, and used to analyse ammonium, nitrite and nitrate. Initial nitrate production rates (using the nitrate evolution over the first 216 h of incubation) were calculated as the slope of a linear regression applied to the nitrate concentrations plotted as a function of time, and were expressed as mg-N h-1 per L of leachate. Three replicates were conducted for each sample. Culturable nitrifiers were quantified at the beginning and the end of tests in all cultures using the MPN (Most Probable Number) method (APHA, 1998). The medium for MPN determination was the same as for the incubation experiment. The tubes were checked for the presence of nitrifying populations, judged by the nitrate formed after 4 weeks. Tubes were scored positive when nitrate was over 40 mg L-1. 2.5. Aerobic nitrite oxidation test Samples of 2.5 ml resuspended leachate solids were added to 500 ml glass bottles containing 100 ml of sterile nitrification medium (per liter: 6 g of CaCO3,, 0.0272 g of KH2PO4, 0.0294 g of CaCl2.2H2O, 0.492 g of MgSO4.7H2O and 0.5 g of NaNO2). The pH of the medium was adjusted to 7.5 and the bottles were incubated using a rotary shaker (150 rpm) at room temperature (28 ºC) for 240 h. During incubation, samples were taken 10 times at different times, and used to analyse nitrite and nitrate. Initial nitrite oxidation and nitrate production rates were calculated as the slope of a linear regression applied to the data from the first 48 hours and expressed as mg-N h-1 per L of leachate. Three replicates were applied for each leachate sample. 2.6. Anaerobic nitrate reduction test Potential nitrate reduction was determined on the basis of the initial nitrate consumption and nitrite production rates in short-term incubations. 2.5 ml inoculum was added to 100 ml glass bottles closed with rubber stoppers. The bottles contained 100 ml of denitrification medium (Alef and Nannipieri, 1995) (per liter: 10 g of sodium acetate, 20 g of KNO3, 0.5 g of

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K2HPO4, 0.2 g of MgSO4.7H2O, 10 ml trace element solution, containing per liter: 1000 mg of Fe(III) citrate, 10 of mg MnCl2.4H2O, 5 mg of ZnCl2, 2.5 mg of KBr, 2.5 mg of KI, 0.005 mg of CuSO4, 1000 mg of CaCl2, 1 mg of Na2.MoO4.2H2O, 5 mg of CoCl2, 0.5 mg of SnCl2.H2O, 0.5 mg of BaCl2, 1 mg of AlCl3, 10 mg of H2BO2, 20 mg of EDTA). The pH of the medium was adjusted to 7.8. Cultures were flushed with nitrogen gas for 5-10 minutes, maintained under anaerobic conditions and incubated at room temperature (28 ºC) for 2 weeks. Eight samples were drawn over time and analysed for nitrite and nitrate as described below. Initial denitrification rates were determined by linear regression on the basis of changes in the nitrite concentrations over the first 48 hours and expressed as mg-N h-1 per L of leachate. Three replicates were performed for each leachate sample. Anaerobic nitrate reducers were enumerated at the start and the end of incubation using a MPN method (APHA, 1998). The medium for MPN was same as for the nitrate reduction test. Tubes were judged as positive if gas formed within 3 weeks. As gas production may not necessarily indicate N2 production, potential anaerobic nitrate reducers in the incubations with S0 dry season sample were isolated for further characterisation by spreading samples on agar plates containing a medium identical to the one used in the denitrification test. Agar plates were subsequently placed into anaerobic jars, flushed (5-10 minutes) with nitrogen gas and incubated at room temperature for 1 week. Colonies were purified by repeated transfers until a pure isolate was obtained. To test the potential of nitrate reduction, five isolates were inoculated into denitrification medium containing 2000 mg-N L-1 of nitrate. Cultures were flushed with nitrogen gas for 5 to 10 minutes, incubated at room temperature for 72 h. Samples were taken at the beginning and at the end and analysed for nitrate and nitrite. 2.7. Anammox tests The potential for anaerobic ammonium oxidation was determined on the basis of simultaneous ammonium and nitrite oxidation rates and the presence of hydrazine as an intermediate compound. 2.5 ml of resuspended leachate solids was added to 100 ml glass bottles, which were closed with rubber stoppers. The bottles contained 100 ml of anammox medium (per liter: 0.76 g of (NH4)2SO4, 0.414 g of NaNO2, 0.20 g of KHCO3, 0.174 g of K2HPO4; 0.075 g of CaCl2, 0.102 g of MgCl2; 2 ml of trace element solution 1, and 1 ml of trace element solution (Egli et al., 2001), set to pH 7.5). Cultures were made anaerobic by flushing them with nitrogen gas for 5-10 minutes, pH increased had increased slightly, to pH 7.7. Subsequently, the cultures were incubated at 37 0C for 192 h. Samples were taken over time and were analysed for ammonium, nitrite, nitrate and hydrazine. Initial anaerobic ammonium oxidation rates was calculated as the slope of a linear regression applied to the ammonium and nitrate concentrations during the first 48 hours, and expressed as mg-N h-1 per L of leachate. Three replicates were applied for each leachate sample. To determine whether anammox activity was occurring under field conditions, and the effect of leachate on anammox activity, additional anammox experiments were done. 2.5 ml inoculum from S0 samples in the wet season was added to 100 ml glass bottles, containing either sterilized, undiluted or ten times diluted leachate. Nitrite was added to a final concentration of 1000 mg NO2

--N per liter for undiluted and 100 mg of NO2--N per liter for 10

times diluted leachate. Cultures were flushed with nitrogen gas for 5-10 minutes, maintained under anaerobic conditions and incubated at 37 ºC for 144 h. Leachate was sterilized by autoclaving at 121 oC for 15 minutes.

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2.8. Analytical methods Hydrazine was determined spectrophotometrically (Tal et al., 2006) by adding 1 ml of p-dimethylaminobenzaldehyde reagent (4 g of p-dimethylaminobenzaldehyde dissolved in 200

ml of methyl alcohol and 20 ml of concentrated HCl) and 0.1 ml concentrated HCl to 5 ml of sample and, after 10 min of incubation at room temperature, determining the absorbance at 458 nm. Standard curves (5 to 200 µg/l hydrazine) were prepared from hydrazine sulphate. Ammonium was analyzed by the colorimetric method of Nessler (APHA, 1998). Nitrite and nitrate were analyzed according to Keeney and Nelson (1982). 2.9. Statistical analysis Data on initial N-arginine mineralization, nitrification, aerobic nitrite oxidation, denitrification and anammox rates were transformed to obtain normal distributions and analyzed by t-tests to establish the impact of season or sampling site. Differences were accepted as significant at the P = 0.05 level. All statistical computations were undertaken using the SPSS programme for Windows, version 12. 3. Results Table 1 presents seasonal variation in leachate quality collected in August 2005 (wet season) and January 2006 (dry season) in the natural (S0) and the artificial (S6) ponds. Ammonium concentrations and Chemical Oxygen Demand (COD) during the dry and wet seasons in S0 were higher than in S6. Acidities of the leachates were relatively similar, while dissolved oxygen (<1.1 mg/l) was only detected during the wet season in S0 and S6. Due to the high chemical leachate background (that is, background concentrations of ammonia in particular), inocula for incubations were prepared from washed pellets obtained after centrifuging the leachates, to avoid the occurrence of nitrogen transforming processes other than the process of interest. 3.1. Arginine mineralization The initial rate of ammonium release from arginine deamination varied to a limited extent between seasons and sampling sites (Table 2). The initial rates of ammonium release were slightly higher in S0 than in S6. The initial rates of aerobic ammonium release were always Table 1. General chemical properties of leachate in a natural (S0) and artificial concrete (S6) treatment pond at the Jatibarang landfill, Semarang, Indonesia, during two seasons (dry and wet). Means are shown with standard deviations. n.d. not detectable (< 0.1 mg L-1). Seasons Parameter Dry (August 2005) Wet (January 2006) S0 S6 S0 S6 pH 8.33 ± 0.06 8.23 ± 0.03 7.88 ± 0.05 8.00 ± 0.03 NH4-N (mg-N L-1) 1255 ± 25 26.2 ± 2.0 1368 ± 38 500 ± 11.6 COD (mg L-1) 5992 ± 30 415 ± 8.4 4465 ± 92 1060 ± 25 Dissolved oxygen (mg L-1) nd nd 0.8 ± 0.04 1.1 ± 0.2

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Table 2. Initial rates of nitrogen transformation process in cultures inoculated with microbial communities from a natural leachate treatment pond (S0) and an artificial concrete pond (S6) at the Jatibarang landfill, Semarang, Indonesia. Parameters Site Season Dry season Wet season Aerobic arginine mineralization Rates of ammonium release (mg-N h-1 per L of Leachate)

S0 S6

28.11 + 7.80 24.39 + 0.20

23.56 + 2.15 21.49 + 2.58

Average 26.25 + 2.63a 22.52 + 1.46a Anaerobic arginine mineralization Rate of ammonium release (mg-N h-1 per L of Leachate)

S0 S6

18.19 + 3.12 10.33 + 1.89

10.33 + 1.89 09.95 + 3.12

Average 14.26 + 5.55a 10.14 + 0.27b Nitrification Rate of nitrate production (mg-N h-1 per L of Leachate)

S0 S6

0.057 + 0.001 0.058 + 0.0007

0.032 + 0.006 0.028 + 0.001

Average 0.058 + 0.0006a 0.030 + 0.003b Aerobic nitrite oxidation Rate of nitrate production (mg-N h-1 per L of Leachate)

S0 S6

1.19 + 0.11 1.20 + 0.09

1.08 + 0.11 1.74 + 0.10

Average 1.19 + 0.07a 1.39 + 0.27a Anaerobic nitrate reduction Rate of nitrite production (mg-N h-1 per L of Leachate)

S0 S6

8.87 + 0.11 8.41 + 0.17

8.22 + 0.32 8.16 + 0.40

Average 8.85 + 0.02a 8.19 + 0.04a Anammox Rate of ammonium oxidation (mg-N h-1 per L of Leachate)

S0 S6

8.48 + 1.03 7.80 + 1.43

4.99 + 0.34 4.05 + 0.54

Average 8.14 + 0.48a 4.52 + 0.66b Rate of nitrite oxidation (mg-N h-1 per L of Leachate)

S0 S6

10.48 + 2.73 18.66 + 0.49

6.66 + 0.021 5.10 + 1.12

Average 14.57 + 5.78a 5.88 + 1.10b All data are presented as means of three replicates. The standard deviation is given behind the mean values. Different letters behind the means indicate significance differences (p < 0.05).

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higher than the rates of anaerobic ammonia release. The initial rates of aerobic ammonium release were not significantly different between seasons. However, the rates of anaerobic ammonium release in samples from dry season (14.28 ± 4.88 mg-N h-1 per L of leachate) was higher than in the wet season (9.71 ± 9.17 mg-N h-1 per L of leachate). Similar results were obtained with mineralization tests using peptone (data not shown).

Fig. 2. Changes in ammonium (○), nitrite (●) and nitrate (Δ) concentrations in nitrifying cultures inoculated with microorganisms from two different leachate ponds (Figure 1) in two seasons: (A) S0 and (B) S6 sampled in the dry season, (C) S0 and (D) S6 sampled in the wet season. The error bar is the standard deviation from the mean of triplicate incubations.

3.2. Nitrification A significant seasonal influence was also observed for initial nitrification rates (t-test, p < 0.05), again revealing a higher rate during the dry season (Table 2). While initial rates of nitrification were low, prolonged incubation revealed a considerable nitrification capacity: nitrate and nitrite concentrations strongly increased and ammonia concentrations declined after 10 to 14 days (Fig. 2). Although ammonia concentrations decreased faster in the incubations with samples from the dry season (fig. 2A-B) than in the wet season (fig. 2C-D), the concentrations of nitrate and in particular nitrite increased clearly only after about 300 hours while for the samples from the wet season the concentrations of nitrite and nitrate started to increase slightly earlier. In all cultures, ammonium was not completely converted into nitrate, and nitrite remained present in the medium. The NO2-N and NO3-N produced at the end of the 22 day long incubation corresponded to 84 to 96% of the consumed ammonium. The numbers of nitrifying microorganisms were comparable between all samples, and the density of culturable nitrifiers increased only slightly during the incubation (Table 3).

0

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Time (h)

NH 4

(mg-

N L-1

)

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and

NO 3

(mg-

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(m

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L-1)

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d N

O3

(mg-

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-1)

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0 48 96 144 192 240 288 336 384 432 480 528

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NH 4

(m

g-N

L-1)

0

20

40

60

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NO

2 an

d N

O3

(mg-

N L

-1)

0

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40

60

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0 48 96 144 192 240 288 336 384 432 480 528

Time (h)

NH 4

(m

g-N

L-1)

0

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40

60

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100

NO

2 a

nd N

O3 (

mg-

N L

-1)

A B

C D

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Table 3. Most Probable Number (MPN) estimates of nitrifying and denitrifying bacteria in cultures inoculated from two leachate treatment ponds (S0, S6) in two different seasons, at the beginning and at the end of the experiment.

Cultures Nitrifiers (Log MPN L-1)

Denitrifiers (Log MPN L-1)

Initial Final Initial Final

S0 dry season

3.97

4.53

3.74

7.62

S6 dry season

3.61

4.21

3.79

7.57

S0 wet season

3.80

4.36

3.83

7.74

S6 wet season

3.99

4.64

3.67

7.15

3.3. Aerobic nitrite oxidation In addition to overall nitrification, the second step of nitrification, nitrite oxidation, was also studied as a separate process (Table 2). The initial potential rate of nitrite oxidation did not differ significantly between seasons, while in contrast the initial rate for the complete nitrification process was lower during the dry season. However, the initial nitrite oxidation rates were 20 to 70 times higher than the overall ammonia nitrification potentials, implying that the potential for aerobic nitrite oxidation exceeded that of aerobic ammonia oxidation. Figure 3 shows how nitrite and nitrate concentrations changed simultaneously. The nitrate concentrations increased steadily until the end of the test, after 240 h. The nitrate concentrations after 240 h in incubations with inocula from the S0 and S6 sites were 37.8 and 34.6 mg-N L-1 respectively, in the dry season, and 48.9 and 67.3 mg-N L-1, respectively, in the wet season. Fifty to 91% of nitrite was converted to nitrate. 3.4. Anaerobic nitrate reduction In anaerobic nitrate reduction tests, the consumption of nitrate and in particular the increase of nitrite was relatively slow for all tested samples, but accelerated after 144 h (Fig. 4 A-B). The recovery of nitrate-N in the produced nitrite-N for the incubations containing S0 and S6 inocula was 90% and 91% respectively for samples in the dry season, and 81% and 79% respectively, for the wet season samples. Thus, nitrate reduction appeared to stop at nitrite. Therefore, initial rates of anaerobic nitrate reduction were calculated from nitrite production. These rates were not influenced by season (t-test, p > 0.05) (Table 2). The numbers of culturable denitrifying microorganisms were initially almost equal in all cultures, and had increased by 4 orders of magnitude by the end of the incubation (Table 3), in line with the rapid increase in nitrite concentrations after 144 h (Fig. 4A-D). Five strains were isolated and

Microbial nitrogen transformation potential in surface run-off leachate from a tropical landfill

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Fig. 3. Changes in nitrite (●) and nitrate (Δ) concentrations in nitrite oxidizing cultures inoculated with leachate from two different treatment ponds, sampled in the dry and wet season: (A) S0, dry season; (B), S6, dry season; (C) S0, wet season and (D) S6, wet season. The error bar is the standard deviation from the mean of three replicate incubations. characterised from incubations inoculated with the S0 dry season sample. The differences in colony morphology and color revealed diversity in the nitrate reducing communities. These isolates did not appear to degrade nitrate beyond nitrite: reduction to nitrite took place during the first 72 h, after that the nitrite concentration in medium remained stable untill the end of the incubation (Table 4). 3.5. Anaerobic ammonium oxidation (anammox) The initial rates of anammox differed significantly between the seasons, with nearly two times higher rates in the dry season compared to the wet season (p < 0.05, Table 2). A simultaneous decrease in ammonium and nitrite concentrations was found in all incubations, with the NH4-N/NO2-N ratio ranging from 0.67 to 1.09 (Figure 5). The concentrations of ammonium and nitrite decreased relatively slowly during the first 24 h, and then decreased faster until 144 h. Concentrations of hydrazine (an intermediate in anammox) peaked between 96 and 120 h (Fig. 6).

0

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0 24 48 72 96 120 144 168 192 216 240

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and

NO 3

(mg-

N L-1

)

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and

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and

NO 3

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)

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0 24 48 72 96 120 144 168 192 216 240

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NO 2

and

NO 3

(mg-

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)

B

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Fig. 4. Changes in nitrate (∆) and nitrite (●) concentrations in anaerobic nitrate reducing cultures inoculated with leachate from treatment ponds S0 and S6 sampled in the dry season (A and B), and S0 and S6 in the wet season (C and D).

Table 4. Nitrite production and nitrate reduction over 72 h by five pure isolates in denitrifying medium containing nitrate (2000 mg-N L-1). WF: white-fluorescent colony, WG: white-grey colony, WM: white-milk colony, Y : yellow, O : orange colony. Variable Pure isolates (codes)

WF WG WM Y O NO2 production (mg L-1)

57.4 ± 32.6

273 ± 21.0

278 ± 48.2

1048 ± 117.1

139 ± 21.7

NO3 reduction (mg L-1)

85.8 ± 10.2 334 ± 14.4 369 ± 12.5 1262 ± 54.7 231 ± 9.6

0

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2500

0 48 96 144 192 240 288

Time (h)

NO

3 (m

g-N

L-1

)

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1400

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(mg-

N L-1

)

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g-N

L-1

)

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L-1)

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L-1

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3 (m

g-N

L-1

)

0

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NO

2 (m

g-N

L-1

)

A

C D

B

Microbial nitrogen transformation potential in surface run-off leachate from a tropical landfill

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Fig. 5. Anammox activity determined by simultaneous disappearance of ammonium (○) and nitrite (●) for cultures inoculated with leachate from two different treatment ponds and two seasons: (A) S0, dry season; (B) S6, dry season; (C) S0, wet season and (D) S6, wet season. The error bar shown is the standard deviation from the mean of three replicate incubations.

Fig. 6. Changes Hydrazine concentrations in anammox cultures inoculated with leachate sampled from: (●) S6, dry season, (○) S0, dry season, (▲) S0, wet season and (∆) S6, wet Season. The error bar shown is the standard deviation from three replicate incubations.

0

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0 24 48 72 96 120 144 168 192

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The potential for anammox, as determined here in synthetic medium, may benefit the development of an alternative cost-effective treatment process for remove inorganic nitrogen. However, high concentrations of organic compounds (Joss et al., 2009) and lack of nitrite and nitrate (Van de Graaf et al., 1995; Trimmer et al., 2005) in the leachate may inhibit anammox activity; other N-transforming processes (e.g. denitrification) might then be more favorable. In order to assess anammox activity in leachate, a new anammox experiment using sterilized leachate in the incubation medium to more closely resembling the field situation, was performed. In ten times diluted leachate, ammonium increased during the first 24 h (Fig. 7B), and then, like nitrite, decreased steadily until nearly nothing was left after 120 h. The NH4-/NO2-N ratio was 0.72, similar to the value in the experiment using synthetic medium. A relatively high hydrazine concentration was found after 48 h (0.28 mg L-1). The initial rate of nitrite consumption was 6.27 ± 0.58 mg-N h-1 per L of leachate, in the range of initial rates measured in synthetic medium without leachate added (Table 2). Figure 7A also shows a small, initial increase of ammonium during the first 72 h in the experiment with undiluted leachate, suggesting some ammonification. Then, ammonium decreased till the end of the test. The nitrite concentration decreased from the beginning till the end of the experiment. The decrease in ammonia and nitrite occurred with a NH4-N/NO2-

Fig. 7. Anammox activity in (A) sterile, ten times diluted landfill leachate and (B) sterile, undiluted landfill leachate medium. Nitrite was added to a final concentration equivalent to the ammonium concentration. Simultaneous consumption of ammonium (○) and nitrite (●) indicates anammox. The error bar shown is the standard deviation of three replicate incubations.

0

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160

0 24 48 72 96 120 144

Time (h)

NH 4

and

NO 2

(mg-

N L-1

)

0

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800

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1400

1600

0 24 48 72 96 120 144

NH 4

and

NO 2

(mg-

N L-1

) B

A

Microbial nitrogen transformation potential in surface run-off leachate from a tropical landfill

75

N ratio of 0.34, indicating a higher nitrite than ammonium consumption. A small amount of hydrazine (< 0.1 mg L-1) was found at the end of test. The initial rate of anaerobic nitrite oxidation could only be estimated in sterile undiluted leachate (182.47 ± 11.94 mg-N h-1 per L of leachate). 4. Discussion This study reconfirmed our previous observations on the chemical composition of Jatibarang landfill leachate, and the impact of seasons, (Mangimbulude et al., 2009) in that it contains high ammonium concentrations throughout the year but is low in nitrate and nitrite. For that reason, the leachate would need to be treated appropriately in order to minimise potential ammonium toxicity (Kim et al., 2006; Clěment and Merlin, 1995; Ernst et al., 1994). Despite the high ammonium concentrations and previously observed effect of season on chemical composition (Mangimbulude et al., 2009), we observed that the leachate possesses the potential for all microbial nitrogen transformations beyond ammonium and that these potentials were not much affected by seasons. Nitrogen transformation potentials in Jatibarang landfill leachate comprised ammonification, nitrification, anaerobic nitrate reduction, and anammox. The highest potential rates were revealed for the anaerobic processes and for ammonification. The potential for ammonification (anaerobically and aerobically) was 300 to 500 times higher than aerobic nitrification. The source of the high ammonification rates is most likely the abundant availability of protein. Municipal solid waste has been reported to contain 3.8 - 4.2% protein (Barlaz et al., 1990). Through ammonification, nitrogen from refuse is mineralized and accumulates as ammonium in leachate (Burton and Watson-Craik, 1998; Jokela and Rintala, 2003). Ammonium was present in high concentrations, while nitrate and nitrite were very low to absent. This correlates with the very low aerobic nitrification potential, compared to the other nitrogen transformation processes. The potential initial rates of nitrification in the dry and the wet season were nearly a thousand times lower than nitrification rates (nitrate production rate) in a landfill leachate treatment system studied by Im et al. (2000). The low potential relates to the anaerobic characteristics of Jatibarang landfill leachate (Mangimbulude et al., 2009), which likely allows only limited growth of aerobic nitrifiers. In particular the first step in nitrification, ammonia oxidation, appears to limit the overall N-mineralisation: the potential rates of the complete nitrification process were more than twenty times lower than the rates of the second step, nitrite oxidation. The products of nitrification, nitrate and nitrite, are essential to allow for full mineralisation of organic nitrogen and ammonia to harmless nitrogen gas, by either denitrification or anammox. The potentials for these anaerobic processes were, despite the low concentrations of nitrate and nitrite, nearly as high as the ammonification rates and much higher than nitrification potential, which supplies the nitrite and nitrate needed for the growth of nitrate reducers and anammox bacteria. The higher potentials suggest relatively high numbers of the latter microorganisms, which may relate to their competitive abilities for nitrite and metabolic versatility to use also other electron acceptors besides nitrate and nitrite. Nitrate reducers are often capable of using other electron acceptors, such as oxygen and iron. Genome sequencing of the anammox bacterium Kuenenia stuttgartiensis followed up with

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physiological experiments suggested that this bacterium is also metabolic versatile in the use of electron donors and acceptors, it is for example capable of reduction of iron with formate as electron donor (Strous et al, 2006). Furthermore, anammox bacteria have a higher affinity for nitrite than nitrifiers (Strous et al., 1999). The simultaneous consumption of ammonium and nitrite plus hydrazine production in the medium was taken as evidence for anammox activity (Egli et al., 2001). The proportion in which ammonium and nitrite were consumed in mineral media was comparable to the ratio generally observed for anammox processes (0.7 to 1.0: Egli et al., 2001). The decrease in ammonia and nitrite in the experiment using landfill leachate as medium occurred with a NH4-N/NO2-N ratio of 0.34, indicating a higher nitrite than ammonium consumption. This may possible be due to simultaneous occurrence of ammonification of leachate components, plus possibly nitrite reduction to ammonium, generating ammonia next to the consumption by the anammox process. The initial rates of anaerobic ammonium oxidation were 2 to 4 times higher than the rate of anaerobic ammonium removal reported by Liang and Liu (2008). The initial potential rates of anaerobic nitrate reduction were almost three times higher than the rates reported by Wu et al. (2009). An intriguing observation in our study is that neither experiments on denitrification potential, nor studies on isolates revealed full denitrification; nitrate reduction appeared to stop at nitrite. Nitrite accumulation is frequently observed during denitrification of waste water and leachate (Wilderer et al., 1987; Martienssen and Schöps, 1997). Nitrite accumulation is influenced by carbon limitation and composition (Almeida et al. 1995, van Rijn et al. 1996). Nitrite reduction is also inhibited by nitrite itself when the concentration is higher than that to which the microorganisms have been acclimated (Chung and Bae, 2002). Overall, the potential of full nitrification in the Jatibarang landfill leachate is the lowest of all nitrogen transformation processes. This suggests that the low level of oxygen in the leachate ponds limits the establishment of active nitrifying communities and their activities. In order to improve the quality of the leachate, and to enable an acceptable concentration of inorganic nitrogen for the discharge of leachate into the river Kreo, implementing a two-step aerobic/anaerobic system needs to be considered to make more optimal use of the current leachate collection system (Fig. 1). This could done by introducing oxygen during the first step in concrete collection pond S5 to a level above 2 mg L-1 (Prinčič et al., 1998) and allowing for further conversion to dinitrogen gas during a second, anaerobic step, in pond S6. Two-steps processes are already applied in nitrogen removal. Biological nitrogen removal through a two-step process of nitrification and denitrification is usually suggested for the treatment of landfill leachate with high ammonium concentrations (Doyle et al., 2001; Klimiuk and Kulikowska, 2004). In the first stage ammonium is converted aerobically to nitrate and in the second stage nitrate is converted to nitrogen gas (Mpenyana et al., 2008). Recently, a single reactor system for Stable High Activity ammonium Removal Over Nitrite reactor system (SHARON), was combined with the anammox process to achieve nitrogen removal (Paredes et al., 2007; Hellinga et al., 1998). The SHARON process favours ammonia oxidisation to nitrite but not further, by employing short retention times that select against nitrite oxidisers. Therewith, the SHARON process diminishes oxygen requirements, which constitute a major fraction of operating costs of biological nitrogen removal processes (Wang et al., 2007). In the combined SHARON-anammox process an ammonium-nitrite mixture is initially produced by aerobically converting 50% of the ammonium to nitrite, which during a subsequent anaerobic stage is converted to dinitrogen gas within a single reactor with a

Microbial nitrogen transformation potential in surface run-off leachate from a tropical landfill

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retention time 1 – 2 days (Schmidt et al., 2003; Paredes et al., 2007). Generally, denitrifying bacteria and anammox bacteria compete for nitrite (Zhang at al, 2010), and denitrifying bacteria may outcompete anammox bacteria (Tal et al., 2006; Kumar and Lin, 2010). High concentrations of organic compounds might inhibit anammox activity (Joss et al., 2009). This indicates that nitrification-denitrification is the most ideal solution for N removal from landfill leachate. However, this option is not applicable for the Jatibarang landfill leachate, as ammonification occurs at a higher rate than ammonium oxidation and anaerobic nitrate reduction did not proceed beyond nitrite. By feeding the second pond, S6, with a mixture of ammonia and nitrate, as the result of prior aerobic nitrification, a combined anaerobic nitrate reduction (to nitrite) and anammox (using nitrite) would allow for the conversion of ammonia to nitrogen gas. Our experiments on anammox activities in leachate indicate that the process will perform under conditions of high COD. Several researchers have reported a successful link between denitrification and anammox in natural environments when the denitrifying bacteria are not competing with anammox bacteria for nitrite (Dalsgaard and Thamdrup, 2002; Dalsgaard et al., 2005; Hulth et al., 2005; Kumar and Lin, 2010). Residence times in the ponds S in the ponds S5 and S6 would need to be sufficiently long enough to allow for complete conversion. Several researchers have reported that the residence time for nitrification in landfill leachate varied from 5 days to 250 days (Haarstad, 1999; Spagni et al., 2008; Mehmood et al., 2009). Based on initial rate of aerobic ammonium oxidation (0.028– 0. 058 mg-N h-1 per L of leachate) and the initial ammonium concentration (ranging between 900 – 1200 mg L-1), it can be calculated that about up to 50% of ammonium will be converted to nitrate if the retention time at S5 would be 400 - 600 days. The current residence time in S5 pond ranges from 2 – 4 days (Mangimbulude et al., 2009). However, when oxygen would be introduced to the S5 pond, it can be expected that the ammonia oxidising communities as well as their activities will increase, as we also observed in our incubation experiments, and thus residence times can be lower. Im et al. (2000) reported that the maximum nitrification rate in their landfill leachate treatment was 20 mg-N –h-1 per L of leachate. At that rate, about 50% of nitrate present in Jatibarang leachate will be removed within one day. The next step is using pond S6 for anaerobic N mineralisation with the remaining ammonium and the nitrate formed during nitrification. Based on initial rates of anaerobic nitrate reduction and anammox (8.16– 8.47 mg-N h-1 per L and 5.10– 18.66 mg-N h-1 per L, respectively), it can be calculated over 80% of nitrate resulting from nitrification will be removed in the S6 pond if the residence time would be 1 - 4 days, which corresponds to the current residence time (Mangimbulude et al., 2009). Fan et al. (2006) reported that most treatment plants have difficulty to meet environmental quality standards due to variation in leachate quantity and composition over seasons. Minor seasonal effects (less than three times differences) were observed on potential nitrogen transformation rates, but seasons were previously found that have a big impact on residence time and concentrations of N compounds (Mangimbulude et al., 2009; this study). Especially the larger leachate quantities result in higher loading of ammonium during the wet season. The hydraulic retention time of the ponds became shorter during the wet season (Mangimbulude et al., 2009), but initial potential rates of nitrification and anammox increased and partially counteract the decreased residence time. In a tropical climate, seasonal effects on leachate quality and quantity are critical factors to be considered in the design and performance of leachate treatment facilities.

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Acknowledgements The authors would like to thank the V. Rutgers fund of the Vrije Universiteit Amsterdam, for supporting the work described here. Thanks go to Mariana, Sylvia Widagdo, Oky Wisnu, Christine Setyaninggrum, Ocon Totoda, and Deny Kurniawan for their assistance in the laboratory and field work. References Alef, K., Nannipieri, P., 1995. Methods in applied soil microbiology and biochemistry.

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4 Microbial community structure and dynamics in

relation to hydrochemistry of surface run-off water from the Jatibarang landfill in Indonesia

Mangimbulude, J.C., Van Straalen, N.M., Röling, W.F.M., 2012

in preparation

Abstract

Microbes play an important role in the natural attenuation of surface run-off water from landfills. This study aimed to gain more insight into the functioning of these communities under the conditions of a tropical landfill at Jatibarang, Indonesia, using four complementary approaches: (1) to determine temporal variation in microbial communities over wet and dry seasons in several years (2004-2010) in the leachate collecting system, using cultivation-independent bacterial community fingerprinting; (2) to establish the relation between community structure and leachate hydrochemistry; (3) to determine the presence of anaerobic degraders of aromatic compounds by amplifying fragments of the bamA gene, encoding an enzyme from the anaerobic benzyol-coA metabolic pathway, and; (4) to determine the actual potential for aerobic and anaerobic phenol degradation by a culturing approach. The diversity in bacterial communities in the period of 2004 to 2006 was higher than in the period from 2006 to 2007. While bacterial communities showed large dynamic changes, leachate hydrochemistry changed little over time. The potential for anaerobic aromatics degradation was present in all years and seasons, throughout the leachate collection system. Restriction profiles of a bamA fragment at two sites in the wet season 2007 were different from other years. The Jatibarang leachate microbes were able to degrade phenol under different redox conditions. We conclude that significant potential for natural attenuation is residing in the local microbial communities, despite the fact that the communities themselves were highly dynamic.

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1. Introduction Landfill leachate is after agricultural activities the second source of contamination of surface water and groundwater (EPA, 1990). Consequently, the remediation of landfill leachate should be considered in environmental protection policies. Monitored natural attenuation (MNA) has received attention as a cheap alternative to techniques that actively decontaminate polluted soil or groundwater (Selberg et al., 2005; Bombach et al., 2010). Natural attenuation is the reduction in mass or concentration of a compound in groundwater or surface water over time or distance from source of constituents of concern due to naturally occurring physical, chemical and biological processes, such as biodegradation, dispersion, dilution, adsorption and volatilization (EPA, 1999). Among these processes, biodegradation is often the primary mechanism for (complete) contaminant removal (Margesin and Schinner, 2001; EPA, 1999). Natural biodegradation processes rely on the collective abilities of microorganisms to metabolise pollutants under prevailing environmental conditions (Zadra et al., 1998). Therefore, gaining insight in the microorganisms present in landfill leachate and their potentials to degrade toxic compounds, such as phenolics, is very important with respect to monitoring, predicting and, when needed, enhancing biodegradation processes (Scow and Hicks, 2005, Röling and van Verseveld, 2002). Monoaromatic compounds are major pollutants in landfill leachate (Kjeldsen and Christopheren, 2001; Reitzel and Ledin, 2002). Phenol was the hazardous aromatic compound reported to occur in highly organic and heterogeneous municipal solid waste (MSW) in a tropical developing country (Swati and Joseph, 2007). Phenol poses a serious threat to environments because at about 1 mg L-1 and higher, it affects aquatic life (Singh et al., 2009). Degradation of phenols may be carried out by microorganisms under aerobic and anaerobic conditions (Boyd et al., 1983; Ronen and Abeliovich, 2000; Chan et al., 2005; Nair et al., 2008). Phenolic compounds are encountered in many landfill leachates, including leachate from the Indonesian Jatibarang landfill, which tends to be anaerobic (Mangimbulude et al., 2009). Anaerobic metabolism of phenol and other aromatic compounds such benzene, toluene, ethylbenzene, and xylenes (BTEX) generally proceeds via the benzoyl-CoA degradation pathway (Kuntze et al., 2011). The evaluation of the potential for natural attenuation of leachate contaminants can be achieved by combining community structure analyses with the detection of functional genes encoding key enzymes of degradation pathways (Nölvak et al., 2012; Bombach et al., 2010). Key enzymes involved in anaerobic degradation of aromatic compounds are benzylsuccinate synthase that initiates the degradation of several aromatic compounds (Bombach et al., 2010) and 6-oxocyclohex-1-ene-1-carbonyl CoA hydrolase that is responsible for ring opening in the benzyol-CoA degradation pathway (Kuntze et al., 2008). Monitoring the genes encoding benzylsuccinate synthase α-subunit (bssA) and 6-oxocyclohex-1-ene-1-carbonyl CoA hydrolase (bamA) revealed a broad range of anaerobic aromatic hydrocarbon-degrading microorganisms in groundwater downgradient of a Dutch landfill (Staats et al., 2011). The differences in environmental conditions affect landfill leachate composition in tropical and temperate regions (Mangimbulude et al., 2009) and may affect microbial community composition and activities. For instance an increase in moisture content in landfill, e.g. due to seasonal rain as occurs in the tropics, enhances the anaerobic degradation processes by facilitating the redistribution of substrate and nutrients and the dispersal of microorganisms, leading to an increase in growth rate (Augenstein and Pacey, 1991). Studies

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on microbial communities and diversity in landfill leachate have so far mainly been done for European and American countries in temperate climate zones (e.g. Barlaz et al.,1989; Fielding et al., 1988; Ludvigsen et al., 1999; Röling et al, 2001, Staats et al., 2011). We previously studied the natural attenuation of landfill components in surface leachate run-off of the Jatibarang landfill, near Semarang, Indonesia (Mangimbulude et al., 2009), with particular emphasis on the microbial transformation potentials of nitrogenous compounds (Mangimbulude et al., 2012). At Jatibarang, the leachate flows from the landfill via a number of leachate collection ponds to a nearby river. In this study, we focused on the microbial communities along the flowpath, with special emphasis on degradation of phenols. The aims were (1) to determine temporal variation in microbial communities over seasons in several years (2004-2010) along the flowpath from landfill to river, using denaturing gradient gel electrophoresis (DGGE) profiles, (2) to relate variation in community structure to leachate hydrochemistry, and (3) to determine the presence of anaerobic aromatic degraders in landfill leachate by targeting bamA genes. The composition of microbial communities and their functional genes was studied using cultivation-independent molecular analyses, as this approach can be done directly without the need for cultivation, which generally only reveals a minority of the present microorganisms (Amann et al., 1995). Furthermore, molecular approaches are more reproducible and faster. However, combining functional gene approaches with culturing of phenol degraders is important to better understand the capabilities of phenol degraders. This culturing comprised the fourth aim of this paper: to determine the actual potential for aerobic and anaerobic phenol degradation by a culturing approach.

2. Material and methods 2.1. Sampling site and time Leachate was collected at five locations at the Jatibarang landfill, Semarang, Indonesia: three locations along the flowpath from landfill to the river Kreo through the leachate collecting system, S0, S1,S6 and two locations in the river Kreo: one is upstream (S9) and one is the meeting point between leachate and river (S8). For site characteristics, see Mangimbulude et al., 2009. In August 2004 and 2005 (dry season), January 2006 (wet season), January and June 2007 (the wet and the dry season), August and November 2008 (the dry and wet season) and in December 2010 (the wet season) leachate was collected in sterile bottles. The bottles were capped and transferred to the laboratory and stored at 4 oC. The next day, 100 ml was vacuum filtered (0.2 μm pore-size filter). The filter was frozen until DNA isolation. 2.2 Aerobic and anaerobic phenol degradation experiments Aerobic biodegradation of phenol was examined by inoculating 2.5 ml of landfill leachate, from the S0 and S6 ponds in the wet and dry season in 2007, into 500 ml glass bottles containing 100 ml sterile mineral medium modified from Boyd et al (1983) (per liter: 0.27 g of KH2PO4, 0.35 g of K2HPO4, 0.53 g of NH4Cl, 0.10 g of MgCl26H2O, 73 mg of CaCl2.2H2O, 20 mg of FeCI2.4H2O, and 1 ml of trace elements). The pH of medium was adjusted to 7.0, bottles were flushed with air for 5 minutes, and closed with rubber stoppers. Phenol solution was added into the culture bottle using a sterile syringe to a final concentration of 10 mg L-1, the inoculated bottles were incubated in triplicate using a rotary

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shaker (150 rpm) at room temperature for 30 days. The bottles were sampled at day 0, 7 and 30, and analysed for phenol and COD. The same medium, phenol concentration and pH were applied for anaerobic experiments, but the volume of culture bottles was 100 ml. The bottles were flushed for 5 to 10 minutes with nitrogen gas, and maintained under anaerobic conditions. To create different redox conditions, nitrate or sulfate solutions were added to culture bottles separately, to a final concentration of 80 mg SO4

2- L-1 (added as Na2SO4) and 400 mg NO3

- L-1 (as KNO3) respectively. Cultured bottles were incubated and analysed as for the aerobic experiments, except that the incubation lasted for 60 days. 2.3. DNA isolation and amplification DNA was isolated using the MoBio Power Soil™ kit (Carlsbad, CA) following manufacturer’s instructions. PCR amplification was conducted in a total volume of 25 µl reaction mixture in each reaction, containing 1 μl of forward (F357) with GC clamp and reverse (R518) primers (Muizer et al., 1993), 1 μl bovine serum albumin, 12.5 1 μl 2X FideliTaqTM PCR Master Mix (USB), 8.5 μl DNase and RNase-free water and 1 μl DNA template. Cycle conditions for the amplification were as follows: initial denaturation 5 min at 94 oC; 35 cycles with each cycle consisting of 94oC for 0.5 min, 54oC for 0.5 min, 72oC for 0.5 min, and a final elongation at 72oC for 8 min. Amplification of the bamA fragment was conducted in 25 μl (total volume) reaction mixture containing 1 μl of forward (oahF) and reverse (oahR) primers (Staats et al., 2011), 1 μl bovine serum albumin, 12.5 1 μl 2X FideliTaqTM PCR Master Mix (USB), 8.5 μl DNase and RNase-free water and 1 μl DNA template. Cycle conditions for the amplification were as follows: initial denaturation 5 min at 94 oC; 35 cycles with each cycle consisting of 94oC for 0.5 min, 55oC for 0.5 min, 72oC for 0.5 min, and a final elongation at 72oC for 8 min. 2.4. 16S rRNA gene-based fingerprinting of bacterial communities DGGE was performed with the Bio-Rad DCode system. The PCR product was loaded onto 1 mm-thick 8% (wt/vol) polyacrylamide (ratio of acrylamide ⁄ bis-acrylamide, 37.5 : 1) in 1x TAE buffer (20 mM Tris, 10 mM acetate, 0.5 mM EDTA, pH 8.0). The polyacrylamide gel had a denaturing gradient ranging from 30% to 55% (where 100% denaturant contains 7 M urea and 40% formamide). Electrophoresis was performed for 3.5 h at 200 V in 1x TAE buffer at a constant 60 oC. Gels were stained with ethidium bromide, visualized under a Vilber Lourmat (TCP-20-M) UV transilluminator and photographed. All DGGE images were converted, analyzed and normalized with Gel Compar II (Applied Maths, Belgium). Similarity values were calculated using Pearson correlation and visualized by UPGMA (Unweighted Paired Group Method with Arithmetic Means) cluster analysis. 2.5. Restriction analysis of bamA fragments 20 µl of PCR products for bamA fragments (0.35 kb) were added into 15 µl digestion mixture consisted of 0.75 µl rsaI, 2.5 µl buffer 10x NEB4 and 11.75 µl RNase-free water. The mixtures were incubated at 37oC for 3 h. After digestion, 30 µl of PCR product fragments were separated by horizontal gel electrophoresis on 3 % agarose at 100 V for 45 minutes. Gels were stained with ethidium bromide, visualized under a Vilber Lourmat (TCP-20-M) UV transilluminator and photographed.

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E

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This seasonal variation needs to be taken into account when interpretating the changes in hydrochemistry over the 6-year investigation period. The trends were studied qualitatively by considering the changes in the dry seasons between 2004 and 2008, and the wet seasons between 2006 and 2010. This revealed that most hydrochemical parameters decreased only slightly over time, if at all. The decreases over the years were minor (less than 10%), compared to seasonal variation. An exception was COD, for which in particular for the S0 pond a decreasing trend was observed. Iron(II) concentrations at S0 were observed to range from 26.9 to 86.5 mg L-1 and sulfate from 53.4 to 186.5 mg L-1. Sulfate and iron(II) tended to decrease along the flow path from S0 to S8 (for sulfate from 186.5 to 7.13 mg L-1 and for iron (II) from 86.5 to 2.29 mg L-1. Sulfate and iron(II) concentrations in the river at S9 were 1.32 mg L-1 and 0.51 mg L-1 respectively. Nitrate and nitrite were never found in S0 and were less than 0.01 mg N L-1 at S1, except in the wet season 2007 when concentrations of 0.14 mg L-1 for nitrate and 0.063 mg L-1 for nitrite were observed. Nitrate and nitrite concentrations at S8 were in the range from 0.04 to 0.77 mg L-1 and 0.012 to 0.53 mg L-1 respectively. Nitrite and nitrite were occasionally detected at S6 with maximum concentrations were 7.85 mg L-1 and 1.22 mg L-1. Also, dissolved oxygen was never detected at S0 and S1. Increases in dissolved oxygen were observed from S1 to S8 (to 5.48 mg L-1). At S9 dissolved oxygen was higher (7.54 mg L-1). These results showed that changes in redox conditions occurred along the flow path, switching from anaerobic to more aerobic conditions. 3.2. Microbial community dynamics 16S rRNA gene based fingerprinting revealed how the bacterial communities in leachate evolved over time at one particular sampling point (Figure 2A) and how communities changed along the flow path from the landfill to the river (Figure 2B). An obvious decrease in diversity and stabilisation in community structure over time was evident for most sampling points. Between the dry season in 2004 and the wet season in 2006 a high diversity was observed for nearly all sampling points (Figure 2B), with in general over 20 bands visible in the DGGE profiles. Community profiles differed between time points, although close inspection of the indicates that community profiles at a particular sampling location did not change completely over time; most bands remained present. However, between the wet season in 2006 and the dry season in 2007 diversity suddenly dropped strongly (about 3 to 6 clearly visible bands), and from that time on little changes in community structure occurred at most sampling points (Figure 2A). The most dominant community members after the wet season 2006 were minor community members (at most) in the samples taken during the wet season of 2006, as their bands cannot be seen. Only at sampling point S8 still considerable variation over time was observed after 2006, while diversity remained low also there. No obvious effect of season on bacterial community structure is evident between 2007 and 2010. Along the flow path from landfill to river shifts in community structure were observed (Figure 2B). Sampling locations S0 and S1 often showed high similarity in community profiles, irrespective of the year and season. The communities at S6 were often also closely resembling those in S0 and S1, but in particular after the dry season in 2007 the DGGE profiles revealed in addition to the bands dominant for S0 and S1 often also a few additional bands. The community profiles in S8 and S9 were in general more different from S1, S0 and S6, even though they shared some bands with those samples.

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Table 1. Overview of observed bands in restriction analyses profiles of bamA PCR fragments obtained from samples taken at S0 to S9, between 2004 and 2010. See also Figure 3. Sampling sites in Seasons and years

Bands a b c d e f g h

Dry-2004 S0 - + + + - - + - S1 - + + + - - + - S6 - + + + - - + - S8 nd nd nd nd nd nd nd nd S9 nd nd nd nd nd nd nd nd Dry-2005 S0 - - - + - - + S1 - + + + - - - - S6 + + - + - + + + S8 + + - + - + + + S9 nd nd nd nd nd nd nd nd Wet-2006 S0 + - - + - + + + S1 nd nd nd nd nd nd nd nd S6 nd nd nd nd nd nd nd nd S8 nd nd nd nd nd nd nd nd S9 - + + + - - - - Dry-2007 S0 - + + + - - + - S1 - + + + + - + - S6 - + + + - - + - S8 - + + + - - + - S9 - + - + - - - - Wet-2007 S0 - + + + - - + - S1 - + + + + - + - S6 + - - - - + - - S8 - - + + - - + - S9 - + + + + - + - Dry-2008 S0 - + + - - - + - S1 - + - + - - + - S6 - + - + - - + - S8 - + - + - - + - S9 - + + + - - + - Wet-2008 S0 - + - + - - + - S1 - + - + - - + - S6 - + - + - - + - S8 - + - + - - + - S9 - + - + - - + - Wet-2010 S0 - + + + - - + - S1 - + + + - - + - S6 - + + + - - + - S8 nd nd nd nd nd nd nd nd S9 - + + + - - + -

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Figure 2. Bacterial 16S rRNA gene-based DGGE profiles (35–55% denaturing gradient) obtained from surface run-off landfill leachate (S0, S1 and S6), leachate-contaminated river water (S8) and upstream river water (S9) samples, taken in the dry and wet seasons between 2004 and 2010. Figures A and B show the same profiles, but organized in different ways: (A) Dynamics in bacterial community profiles over time, profiles are organized per sampling location, going from S0 to S9. (B) Dynamics in bacterial community profiles along the flowpath, profiles are organized per sampling occasion, from 2004 to 2010.

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3.3. Anaerobic aromatic degrading bacteria in landfill leachate Correctly sized bamA PCR fragments (0.35 kb) were obtained for all leachate samples, and also for river water (S9), suggesting that bacteria with the anaerobic benzoyl-coA pathway were always present, irrespective of year, season and sampling location. In contrast, no correctly sized bssA PCR fragments could be obtained. To assess the diversity and changes in aromatic compound degrading populations, restriction analysis on the bamA fragments was conducted. The results showed that restriction profiles (examples shown in Figure 3) were rather simple, revealing just a few bands and indicating little diversity within samples. Profiles in Figure 3 revealed two (S8) to five (S1) different bands. The total length of the different bands added up to about 0.7 kb maximally for e.g. S1, suggesting two different types of bamA genes being present in that sample (Figure 3). The difference in restriction profiles indicates that different types of anaerobic aromatics degrading bacteria are present. Between sampling points and over time, restriction profiles were rather similar (Table 1). Only some variation was observed at S6 and S8, between 2004 and 2006.

Figure 3. Example of restriction analysis profiles of bamA PCR fragments. Shown are profiles obtained from sampling points (S0, S1, S6, S8 and S9) in the wet season of 2007, L is a 100 bp DNA ladder, + is a positive control. Letters indicate different bands, corresponding to Table 1. 3.4. The potential for phenol degradation Laboratory batch experiments were carried out to evaluate the potential of leachate microorganisms to degrade phenol (10 mg L-1) under different redox-conditions. Within the 30-days incubation period, removal efficiency of phenol in S0 and S6 dry season cultures was

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83% and 77% respectively, while in S0 and S6 wet season culture it was 90% and 95% respectively. Under anaerobic conditions (denitrification and sulfate reducing conditions), the efficiency of phenol removal was lower (less than 50%) than under aerobic conditions. Up to 70% efficiency was reached when incubation was extended to 60 days. Overall phenol removal under denitrifying and sulfate-reducing conditions in the wet season were higher than in the dry season (p<0.05). 4. Discussion The hydrochemistry of surface leachate of the Jatibarang landfill in Semarang, Indonesia, was monitored over six years. Sampling time relative to season (dry or wet) had a significant effect on the hydrochemistry, and the seasonal changes appeared to be larger than the changes over the years. In general, average concentrations of EC, COD, chloride and COD were higher in the dry season than in the wet season, while ammonium was lower. Compared to our previous studies on this landfill, in which temporal dynamics were investigated over a shorter period, but at higher sampling frequency (Mangimbulude et al., 2009, 2012), similar seasonal trends were observed. Also the decrease in hydrochemical parameters along the flow path from landfill to river were similar, and suggested again that dilution is the main attenuating factor (Mangimbulude et al., 2009). During the six years of investigation, and correcting for seasonal effects, the decreases in COD and ammonium concentrations at S0 and S1 were limited, suggestion that over time the contribution of biodegradation of organic matter and oxidation ammonium to leachate attenuation remained low under the prevalent oxygen limited conditions. This is in contrast with our previous study (Mangimbulude et al., 2009) reporting that in a long-term laboratory experiment (700 days) strong decreases in organic leachate and ammonia concentrations in Jatibarang leachate under limited oxygen conditions were possible; this suggests that over time also an increase in biodegradation rates might have been expected under natural conditions. Mangimbulude et al. (2012) reported that the initial potential rate of aerobic nitrification in Jatibarang landfill leachate was lower than other initial rates of microbial nitrogen transforming processes (ammonification, denitrification, anammox). Possibly, the availability of oxygen in the laboratory experiment was higher than under natural conditions, and thereby allowed for higher rates of degradation of ammonia and organic matter in that experiment. Since natural attenuation is a viable option for remediation pollution by landfill leachate, several studies have been performed in order to understand geo-hydrochemistry processes and the microbial communities present in contaminated water and/or leachate. (e.g. Christensen et al., 2001, Röling et al., 2001, Huang et al., 2005). Knowledge on microbial community structure, their capabilities and their effects on the environment, and vice versa, is required to develop tools for predicting, monitoring and, if needed, enhancement of natural degradation in landfill leachate polluted water (Röling et al., 2001). Microbial communities can be assessed by classical culturing approaches. However, these techniques do not reveal the actual microbial community, as only a minor part of microbes are cultivable under laboratory conditions. One possible method to address this problem is to use molecular biology approaches (Amann et al., 1995). Spatial and temporal dynamics of microbial community structure in leachate and leachate-polluted water have been studied

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using molecular approaches (e.g. Lee et al., 2004 Röling et al., 2001; Huang et al., 2005; Brad, 2007; Tian et al., 2005; Staats et al., 2011). In this study obvious changes in microbial community profiles were observed over time, despite the relatively minor changes in hydrochemistry. Thus, it seems that changes in microbial communities were not directly associated with leachate hydrochemistry over time. Complex communities were observed between 2004 to 2006, while much simpler community profiles were revealed after 2006, while hydrochemistry remained rather stable over the 6 y period. Although clear differences in hydrochemistry over seasons were observed, this was not reflected in the community profiles. Thus, a decrease in diversity occurred over time, but this did not affect functioning of the microbial communities, in the sense that concentrations of organic matter and ammonia did not go significantly up or down. Fernandez et al. (1999) also reported that dynamic microbial communities in a waste water treating reactor could sustain stable ecosystem functioning. In contrast, Li et al. (2012) explored microbial community structure in managed aquifer recharge systems and observed that organic concentrations were negatively correlated with diversity at both laboratory and field scales, while McGrady-Steed et al. (1997) reported that microbial functioning in aquatic ecosystems became more stable with increasing diversity. Furthermore, Nanqi et al. (2006) demonstrated that changes in sulfate reducing communities in waste water treatment plants took place along with decreasing alkalinity concentration (compared to the seasonal changes in alkalinity in this study). Succession of microbial communities occurred along the flow path, likely due to changing environmental conditions, in particular with regard to oxygen. Leachate at S0 and S1 tended to be anaerobic (dissolved oxygen concentration was 0 mg L-1), while S6 tended to be semi-aerobic (dissolved oxygen concentrations ranged from 0.5 to 1 mg L-1) while S8 and S9 were aerobic (dissolved oxygen concentrations up to 2 mg L-1). Availability of sulfate in the leachate ponds indicates the potential for sulfate-reducing bacteria to grow in the leachate since other redox components such as NO3, and (insoluble) Fe(III) are generally absent in leachate. Sulfate reducing microorganisms are often observed in surface run-off water (Snow et al., 2008., King et al., 2002; Ogg and Patel, 2011). Anaerobic leachate also often contains toxic aromatic compounds (Christensen et al., 2001; Carmona et al., 2009; Varank, 2011). Therefore, a better understanding of their anaerobic biodegradation is also essential in relation to site clean-up. Many aromatic hydrocarbons degrade very slowly or not at all under anaerobic conditions (Wilson and Bouwer, 1997). Widdel et al. (2010) reported that anaerobic microorganisms utilizing hydrocarbons always exhibit much slower growth than their aerobic counterparts. Consequently, longer time is needed to achieve complete degradation. The mineralization of aromatic compounds by facultative or obligate anaerobic bacteria can be coupled to anaerobic respiration with a variety of electron acceptors, e.g., nitrate, sulfate, iron(III), manganese(II), and selenate (Carmona et al., 2009; Christensen et al., 2001). Carmona et al. (2009) reported that fermentative bacteria can also use aromatic compounds, but usually complete biodegradation only becomes energetically feasible when fermentative microorganisms are accompanied by methanogens or sulfate reducers, with sulfate being more energetically favorable. Anaerobic biodegradation might be considered as a strategy for natural attenuation. In the case of Jatibarang leachate anaerobic degradation of aromatic compounds could possibly take place with sulfate, since this electron acceptor is abundant. Indeed, culturing revealed the potential for anaerobic phenol degradation with sulfate (and also nitrate) as an

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electron-acceptor. Phenol degradation under denitrifying and sulfate reducing conditions can be carried out by different microbial communities (Monserrate and Häggblom, 1997; Shinoda, et al., 2000; Thomas et al., 2002; Qiu et al., 2008). The presence of anaerobic BTEX degrading microorganisms can nowadays also quickly be established by targeting functional genes specifically involved in degradation of monoaromatic hydrocarbons (Staats et al., 2011). The bamA gene, encoding an enzyme involved in the anaerobic degradation of the central intermediate benzoyl-CoA in the metabolism of aromatics, provides a new and widely applicable tool for the detection of all types of anaerobic bacteria capable of degrading a wide variety of aromatic compounds (Kuntze et al., 2008). Using this functional gene, in combination with restriction analysis, rather stable and low-diversity anaerobic aromatic degrading communities over time and space were revealed in Jatibarang landfill leachate and river water. In contrast, a study on a Dutch aquifer polluted with landfill leachate revealed much higher heterogeneity and diversity in bamA fragments over time and space (Staats et al., 2011). Concluding, the changes in microbial diversity in the Jatibarang landfill leachate could not be associated with variation in leachate hydrochemistry. Fast molecular methods revealed a specific, low-diversity community of anaerobic aromatic degraders to be present in Jatibarang landfill leachate, and the potential for anaerobic aromatic degradation was confirmed by more time-consuming cultivation-based methods, which, depending on the source of aromatics, can take more than a year (Botton et al., 2007). This indicates the potential of the occurrence of natural attenuation of Jatibarang landfill leachate under anaerobic conditions. References Amann, R.I., Ludwig, W., Schleifer, K.H., 1995. Phylogenetic identification and in situ

detection of individual microbial cells without cultivation. Microbiological Review 59, 143-169.

APHA, 1998. Standard Methods for Examination of Water and Wastewater, 19th ed. American Public Health Association, Washington, DC.

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Botton, S., Van Harmelen, M., Braster, M., Parsons, J.R., Röling, W.F.M., 2007. Dominance of Geobacteraceae in BTX degrading enrichments from an iron-reducing aquifer. FEMS Microbiology Ecology 62,118-130.

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Brad, T., 2007. Subsurface landfill leachate home to complex and dynamic eukaryotic communities. Ph.D Thesis. Vrije Universiteit Amsterdam, The Netherlands.

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Christensen, T.H., Kjeldsen, P., Bjerg, P.L., Jensen, D.L., Christensen, J.B., Baun, A., Albrechtsen, H.-J., Heron, G., 2001. Biogeochemistry of landfill leachate plumes, Applied Geochemistry 16, 659-718.

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EPA., 1999. Use of monitored of natural attenuation at superfund, RCRA corrective action, and underground storage tank sites.

Fernandez A., Huang, S., Seston, S., Xing, J., Hickey, R., Criddle, C., Tiedje, J., 1999. How stable is stable? Function versus community composition. Applied and Environmental Microbiology 65, 3697-3704.

Fielding, E.R., Archer, D.B., de Macario, E.C., Macario, A.J.L., 1988. Isolation and characterization of methanogenic bacteria from landfills. Applied and Environmental Microbiology 54, 835-836.

Harwood, C.S., Parales, R.E., 1996. The β-ketoadipate pathway and the biology of self-identity. Annual Review Microbiology 50, 553-590.

Huang, L.N., Zhu, S., Zhou, H., Qu, L.H., 2005. Molecular phylogenetic diversity of bacteria associated with the leachate of a closed municipal solid waste landfill. FEMS Microbiology Letters 242, 297-303.

King, J.K., Harmon, S.M., Fu, T.T., Gladden, J.B., 2002. Mercury removal, methylmercury formation, and sulfate-reducing bacteria profiles in wetland mesocosms. Chemosphere 46, 859-870.

Kjeldsen, P., Christophersen, M., 2001. Composition of leachate from old landfills in Denmark. Waste Management Research 19: 249-259.

Kuntze, K., Shinoda, Y., Moutakki, H, Mcinerney, M.J., Vogt, C., Richnow, H.H., Boll, M., 2008. 6-Oxocyclohex-1-ene-1-carbonyl-coenzyme A hydrolases from obligately anaerobic bacteria: characterization and identification of its gene as a functional marker for aromatic compounds degrading anaerobes. Environmental Microbiology 10, 1547-1556.

Kuntze, K., Vogt, C., Richnow, H.H., Boll, M,. 2011. Combined application of PCR-based functional assays for the detection of aromatic-compound degrading anaerobes. Applied and Environmental Microbiology 77, 5056-5061.

Lee, S.-H.., Kim, H.-G., Park, J.B., Park, Y.K., 2004. Denaturing gradient gel electrophoresis analysis of bacterial populations in 5-stage biological nutrient removal process with step feed system for wastewater treatment. The Journal of Microbiology 42, 1-8.

Li, Y.N., Porter, A.W, Mumford, A., Zhao, X.H., Young, L.Y., 2012. Bacterial community structure and bamA gene diversity in anaerobic degradation of toluene and benzoate under denitrifying conditions. Journal of Applied Microbiology 112, 269-79.

Lin, B., Braster, M., Van Breukelen, B.M., Van Verseveld, H.W., Westerhoff, H.V., Röling, W.F.M. 2005. Geobacteraceae community composition is related to hydrochemistry and biodegradation in an iron-reducing aquifer polluted by a neighboring landfill. Applied and Environmental Microbiology 71, 5983-5991.

Ludvigsen, L., Albrechtsen, H.-J., Reingelberg, D.B., Ekelund, F., Christensen, T.H., 1999. Distribution and composition of microbial population in a landfill leachate contaminated aquifer (Grindsted, Denmark). Microbial Ecology 37, 197-207.

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Mangimbulude, J.C.,Van Breukelen, B.M., Krave, A.S., Van Straalen, N.M., Röling, W.F.M., 2009. Seasonal dynamics in leachate hydrochemistry and natural attenuation in surface run-off water from a tropical landfill. Waste Management 29, 829-838.

Mangimbulude, J.C., Van Straalen, N.M., Röling, W.F.M., 2012. Microbial nitrogen transformation potential in surface run-off leachate from a tropical landfill. Waste Management 32, 77-87.

Margesin, R., Schinner, F., 2001. Biodegradation and bioremediation of hydrocarbon in extreme environments. Applied and Environmental Microbiology 56, 650-663.

Monserrate, E., Häggblom, M.M., 1997. Dehalogenation and biodegradation of brominated phenols and benzoic acids under iron-reducing, sulfidogenic, and methanogenic conditions. Applied and Environmental Microbiology 63, 3911-3915.

McGrady-Steed, J., Harris, P.M,, Morin, P.J., 1997. Biodiversity regulates ecosystem predictability. Nature 390, 162–165.

Nair, C.I., Jayachandran, K., Shashidhar, S., 2008. Biodegradation of phenol. African Journal of Biotechnology 7, 4951-4958.

Nanqi, R., Yangguo, Z., Aijie, W., Chongyang, G., Huaixiang,S., Yiwei, L., WAN Chunli, W., 2006. The effect of decreasing alkalinity on microbial community dynamics in a sulfate-reducing bioreactor as analyzed by PCR-SSCP. Science in China Series C: Life Sciences 49, 370-378.

Nölvak, H., Sildvee, T., Kriipsalu, M,. Truu, J., 2012. Application of microbial community profiling and functional gene detection for assessment of natural attenuation of petroleum hydrocarbon in boreal subsurface. Boreal Environmental Research 17, 113-127.

Ogg, A.D., Patel, B.K.C., 2011. Desulfotomaculum varum sp. nov., a moderately thermophilic sulfate-reducing bacterium isolated from a microbial mat colonizing a Great Artesian Basin bore well run off channel. Biotechnology 1, 139-149.

Qiu, Y.L., Hanada, S., Ohashi, A., Harada, H., Kamagata., Y Sekiguchi., Y 2008. Syntrophorhabdus aromaticivorans gen. nov., sp. nov., the first cultured anaerobe capable of degrading phenol to acetate in obligate syntrophic associations with a hydrogenotrophic methanogen. Applied and Environmental Microbiology 74, 2051–2058.

Reitzel, L. A., Ledin, A., 2002. Determination of phenols in landfill leachate contaminated groundwaters by solid-phase extraction. Journal of Chromatography A 927, 175-182.

Röling, W.F.M., Van Breukelen, B.M., Braster, M., Goeltom, M.T., Groen, J., Van Verseveld, H.W., 2000. Analysis of microbial communities in a landfill leachate polluted aquifer using a new method for anaerobic physiological profiling and 16S rDNA based fingerprinting. Microbial Ecology 40, 177-188.

Röling, W.F.M., Van Verseveld, H.W., 2002. Natural attenuation: What does the subsurface have in store?. Biodegradation 13, 53-64.

Röling, W.F.M., Van Breukelen, B.M., Braster, M., Lin, B., Van Verseveld, H.W., 2001. Relationships between microbial community structure and hydrochemistry in a landfill leachate-polluted aquifer. Applied and Environmental Microbiology 67, 4619-4629.

Ronen, Z., Abeliovich, A., 2000. Anaerobic-aerobic process for microbial degradation of tetrabromobisphenol A. Applied and Environmental Microbiology 66, 2372-2377.

Scow, K.M., Hicks, K.A., 2005. Natural attenuation and enhanced bioremediation of organic contaminants in groundwater. Current Opinion in Biotechnology 16, 246-253.

Selberg, A., Viik, M., Peet, K., Tenno, T., 2005. Characteristics and natural attenuation of Pääsüla landfill leachate. Proceeding of the Estonian Academy of Sciences. Chemistry 54, 35-44.

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Singh, S., Singh, B.B., Chandra, R., 2009. Biodegradation of phenol in batch culture by pure and mixed strains of Paenibacillus sp. and Bacillus cereus. Polish Journal of Microbiology 58, 319-325.

Shinoda, Y., Sakai, Y., Ué, M., Hiraishi, A., Kato, N., 2000. Isolation and characterization of a new denitrifying spirillum capable of anaerobic degradation of phenol. Applied and Environmental Microbiology 66, 1286-1291.

Snow, A., Ghaly, A.E., Cote, R., 2008. Treatment of stormwater runoff and landfill leachates using a surface flow constructed wetland. American Journal of Environmental Sciences 4, 164-172.

Staats, M., Braster, M., Röling, W.F.M., 2011. Molecular diversity and distribution of aromatics hydrocarbon-degrading anaerobes across a landfill plume. Environmental Microbiology 13, 1216-1227.

Swati, M., Joseph, K., 2007. Fate of phenol, diethyl phthalate and dibutyl phthalate in bioreactor landfill and controlled dump lysimeters. Proceedings of the International Conference on Sustainable Solid Waste Management, Chennai, India.

Tian, Y.–J., Yang, H., Wu, X.-J., Li, D.–T., 2005. Molecular analysis of microbial community in a groundwater sample polluted by landfill leachate and seawater .Journal of Zhejiang University Science B 6, 165-170.

Thomas, S., Sarfraz, S., Mishra, L.C., Lyengar, L., 2002. Degradation of phenol compounds by a defined denitrifying bacteria culture. World Journal of Microbiology and Biotechnology 18, 57-63.

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5 Hydrochemical characterization of a tropical,

coastal aquifer affected by landfill leachate and seawater intrusion

Mangimbulude, J.C., Goeltom, I., Van Breukelen, B.M., Van Straalen, N.M., Röling, W.F.M.

submitted

Abstract The hydrochemistry of landfill leachate and groundwater is affected by waste degradation processes, but also by external factors such as the geography of the landfilling site. Knowledge on the fate of landfill leachate in tropical countries will be beneficial for monitoring and regulatory purposes. We studied the Keputih landfill close to the sea at Surabaya, Indonesia: (1) to assess leachate and groundwater hydrochemistry with respect to contamination and seawater intrusion, (2) to investigate the seasonal effects on hydrochemical composition; and (3) to determine redox conditions in order to evaluate the potential for natural attenuation through microbe-mediated electron-accepting processes. We document an influence from sea water intrusion on groundwater hydrochemistry on top of the influences from the landfill itself. Leachate had a high electrical conductivity and high COD, and contained high concentrations of NH4

+, HCO3-, SO4

2-, Fe2+ and Cl-. Concentrations were significantly influenced by season, except for COD and SO4

2-. The groundwater at locations surrounding the landfill was also contaminated by leachate and concentrations of groundwater contaminants were higher than national regulatory standards in Indonesia for drinking water. The abundance of SO4

2- in groundwater indicates a large potential for anaerobic biodegradation of organic compounds. Based on the relative concentrations of Cl- and SO4

2- an influence of the sea water on groundwater hydrochemistry was obvious. Landfilling in developing countries often occurs in coastal areas, therefore we emphasize the need to study microbial community structure and functioning in relation to degradation of landfill leachate in tropical coastal areas impacted by seawater infiltration. Keywords: landfill, seawater intrusion, subsurface, natural attenuation

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1. Introduction Landfilling of solid waste is considered one of the better disposal options in developing countries because it is generally more economical than alternatives such as incineration and composting (Johannessen and Boyer 1999). With the increasing generation of solid waste caused by rising urban populations, industrialization and unbalanced socio-economical growth, contamination of the environment has become an important urban issue. Excessive rainwater percolating through a landfill can potentially contaminate the underlying subsurface and groundwater. However, natural processes in the subsurface can remove contaminants derived from landfill leachate. The combined effects of naturally occurring processes (dispersion, dilution, adsorption, volatilization, ion-exchange, precipitation, and biotransformation) will result in the attenuation of pollution: a decrease in concentrations away from the source (Christensen et al. 2001). Biological transformation is the only Natural Attenuation (NA) process that can lead to a complete removal of pollutants. In general, leachate contains high concentrations of reduced components, especially organic matter, CH4 and NH4

+ (Christensen et al. 2001; Kjeldsen et al. 2002). Subsurface microorganisms will oxidize many organic components and simultaneously reduce electron acceptors such as O2, NO3

-, Mn(IV), Fe(III), SO42-, and CO2,

leading to the formation of redox zones in the plume of pollution (Christensen et al. 2001; Kjeldsen et al. 2002). Natural attenuation is a cost-effective and environmentally friendly remediation strategy for contaminated groundwater, but monitoring is required to judge if it is occurring satisfactory, hence: monitored natural attenuation (MNA). The large size of landfills, location(s) of possible contaminant leaks, history of waste dumping, heterogeneous composition of waste, physico-chemical properties, and subsurface hydrogeological complexities of the landfilling area are all factors that have a great influence on monitoring and evaluating attenuation of contaminant plumes (Jorstad et al. 2004; Wilson et al. 2004). Standard practice for monitoring landfill leachate is to lay out a suite of wells for groundwater sampling. Successful MNA on polluted aquifers, including aquifers polluted by landfill leachate, has been reported for temperate regions like the U.S.A. and European countries (Baun et al. 2003; Cozzarelli et al. 1999; Van Breukelen et al. 2003). Especially in tropical countries, MNA of aquifers contaminated by landfills might be an economical and environment-friendly strategy since local authorities generally have few opportunities to implement and maintain intensive treatment facilities. The high tropical temperatures and humidity may possibly even lead to higher rates of biodegradation of contaminants compared to temperate environments (Chaillan et al. 2004). On the other hand, attenuation of landfill contaminants in many tropical countries may be seasonally variable with strong leaching during the wet season and retention of leachate during the dry season (Mangimbulude et al. 2009; Tränkler et al. 2005). These external factors might either positively or negatively affect the activities and distribution of subsurface, pollution degrading microbial communities over time and space. We characterized the seasonal variation in the hydrochemistry of the aquifer impacted by the Keputih landfill, one of the landfills serving Surabaya, Indonesia. This coastal city is the second biggest city in Indonesia, having a population of 2.9 million. This study aimed to (i) assess the heterogeneity in leachate and groundwater quality with respect to contamination by landfill leachate and possible seawater intrusion, (ii) investigate the seasonal effects on

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hydrochemical composition, and (iii) determine the redox conditions in order to evaluate the potential for natural attenuation via terminal electron accepting processes. 2. Materials and Methods 2.1. Field site description The Keputih landfill is situated in a region with a tropical climate (latitude 7o18.20’-7o19.20’and longitude 112o47.60-112o48.60). Generally, the dry period is from May to November, and the wet period runs from December to April. The temperature in the area varies from 22 to 33.5oC and the difference in temperature between the dry and wet season is fairly small (3-5 oC). Average annual rainfall between 1996 and 2006 ranged from 1860 mm to 2000 mm per year (Schroeder et al. 2004; Surabaya Meteorological and Geophysical Agency 2007). The landfill is located in the Eastern part of the greater Surabaya area, occupies about 24 hectares and is surrounded by traditional settlements (Fig. 1). The landfilling area was originally part of a natural salt marsh with a typical coastal wetland vegetation (e.g. mangrove and saltmarsh grasses) (Rachmansyah 2001) and this salt marsh area is extending eastwards to the coast, which is 4 km east of the landfill. The hydrology is strongly influenced by tidal cycles. The sediment of the underlying aquifer is composed mainly of fine sands, silts, and clay with high organic matter content as a consequence of alluvial deposits from the river Brantas. The depth to the base of the aquifer is about 39 m (Rachmansyah 2001). Geologically, the area is situated in the randu glatung basin and a part of pamekasan formation, consisting mainly of alluvial sediments (Van Bemmelen 1949). The local aquifer is shallow, unconfined, and dates back to the Meocene-Pleistocene age. The area has an elevation of 3-4 m above sea level and the groundwater table is at 0 to 4 m below ground surface, depending on the season. The direction of regional groundwater flow is from West to East (Rachmansyah 2001). Landfilling began in the early 1980s and ended in 2001. The landfill received about 4000 m3 of solid waste daily including municipal waste, demolition waste, and industrial waste that is potentially hazardous (Endah and Pudjiastuti 1995). Initially. refuses were dumped in the marsh and covered by soil and sand, but over time the system became an open dump due to lack of a soil cover. At an early stage of the landfilling, refuse was buried in the Western, Northwestern and Northeastern areas (Fig. 1). Over time, dumping was extended eastward until the whole area within the borders was completely covered. At the time of closure, in 2001, the refuse at the Eastern part reached a height of 10 m above ground level. 2.2. Installation of monitoring wells To characterize contamination status and redox processes, monitoring wells (PVC pipe, 5 cm Ø, screen length 50 cm or steel, 3 cm Ø) were installed in and surrounding the landfill (Fig. 1). Between November 2002 and April 2005, a total of 51 wells were installed to a maximum depth of 6 m, and wells encountered the water table within a few metres below the surface. Wells were grouped into nine clusters (Fig. 1) according to their orientation to the landfill. Wells LF1-8 sampled water from the refuse itself, while monitoring wells belonging cluster west (W1-8), northwest (NW1-3), north (N1-12), northeast (NE1-4), east (E1-9), southeast

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(SE1), south (S1-4) sampled groundwater. Reference (Ref1-2) wells were installed ~600 m southwest of the landfill.

Fig.1. Map of the Keputih landfill site near the coast at Surabaya, Java, Indonesia. Monitoring wells (●) were organized in clusters in (LF) and around the landfill: W, western; NW, northern west; N, northern; E, Eastern; S, southern; SE, southeastern. ○ are reference monitoring wells (REF). Numbers indicate well numbers. 2.3. Sampling from monitoring wells and hydrochemical analysis The wells were sampled up to six times using a peristaltic pump, in the dry season (November 2002, July 2003, and November 2003) and wet season (January 2004, April 2004, and April 2005). Temperature, electrical conductivity, dissolved oxygen, and pH were measured using electrodes (HACH, Loveland, Co; WTW, Hamburg, Germany) placed in flow cells. Groundwater samples for analysis of cations NH4

+, Fe2+, Mn2+, were filtered on site (0.1 µm Ø, Sartorius) and transfered to 100 ml PE bottles containing preservative (2% [v/v] concentrated HNO3 for acidification to pH 2). Unfiltered groundwater samples for analysis of chemical oxygen demand (COD) and anions HCO3

-, PO43-, Cl-, NO3

-, and NO2- were kept in

100 ml PE bottles and stored at 4oC until laboratory analysis. Groundwater samples for CH4 and H2S analysis were placed in 30 ml amber serum vials without headspace, capped and kept cool (4 to 10oC). Spectrophotometric analysis was conducted for COD, Cl-, PO4

3-, NO3-, NO2

-, SO4

2- and Mn2+ (using commercially HACH kits (Loveland, Co)), NH4+ (Nesslerisation,

APHA, 1998), Fe2+ (Viollier et al. 2000) and H2S (Cline 1969), while alkalinity was determined by Gran titration (Stumm and Morgan 1981). For CH4 analysis, ten milliliters of groundwater was evacuated using a gas-tight syringe from sampling bottles to a 12 ml crimp-

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capped serum vial, and from the 2 ml headspace volume of the evacuated vial 100 μL of gas was taken and injected into a gas chromatograph (Hewlett-Packard) equipped with flame ionization detector (GC-FID). 2.4. Statistical analysis To compare statistical means of hydrochemical parameters within monitoring well clusters, analysis of variance (ANOVA) following by least-square difference (LSD) post-hoc test were applied to the log-transformed data. All statistical analysis was done using SPSS software version 12. 2.5. Calculation of the seawater fraction in groundwater samples The fraction of sea water in groundwater samples was calculated as follows, under the assumption of conservative mixing (Appelo and Postma 2005): where fsea, is the fraction of sea water in a sample, mcl

-,sample is the chloride concentration of

the sample, mcl-,sea is the chloride concentration in sea water (566 mM), and mcl

-,fresh is the

local background (non-seawater influenced) concentration of chloride, estimated from the groundwater reference well (minimum concentration of 0.4 mM). 3. Results 3.1. Groundwater hydrochemistry The hydrochemistry of water samples taken from the landfill was clearly different from those of samples obtained from the reference wells (Table 1). Concentrations of compounds expected to increase due to landfilling were strongly enhanced: the maximum concentrations of COD and NH4

+ in the landfill (5016 mg/L and 20.55 mM, respectively) were much higher compared to the references (193 mg/L and 0.35 mM, respectively) (Table 1). Microorganisms contribute to changes in hydrochemistry by metabolic conversions of landfill leachate compounds and may produce alkalinity (HCO3

-), Fe2+, CH4 and H2S (Christensen et al. 2001). Alkalinity (HCO3

- concentration) was strongly increased, with maximally 590 mM in the landfill samples, compared to 6.5 mM in the reference wells (Table 1). High concentrations of CH4 (up to 2.3 mM) were observed in the leachate (Table 1). Concentrations of Fe2+ and H2S were not much different between reference wells and wells in the landfill. PO4

3- is mainly derived from refuse decomposition, and this can be seen by a consistently higher concentration of this nutrient in the leachate; a maximum concentration of 0.38 mM was observed in the landfill. The observed wide range in the measured concentrations for the monitoring wells installed in the landfill, indicates considerable spatial heterogeneity. In order to study the spread of leachate into groundwater, the hydrochemistry of groundwater surrounding the landfill was compared to groundwater in reference wells and in the landfill (Table 1). Groundwater in almost all monitoring wells was colored (yellow) indicative of groundwater mixed with landfill leachate (brown). Indicators of the presence of landfill leachate (NH4

+, COD and alkalinity) were also found to be enhanced in wells all

freshclseacl

freshclsampleclsea mm

mmf

,,

,,

−−

−−

−

−=

106

Table 1. Hydrochemistry of groundwater extracted from different monitoring well clusters in and around the landfill during the wet and the dry season West North-West North North-East East S-E South Landfill Reference

Wet Dry Wet Wet Dry Wet Dry Wet Dry Wet Wet Wet Dry Wet Dry

Temperature

pH

EC

COD

Ammonium

Nitrate

Nitrite

Iron (II)

Manganese (II)

Sulfate

Sulfide

Methane

Alkalinity

Phosphate

Chloride

23-27

6.9-7.5

2.2-9.8

81-657

0.6-3.3

0-0.5

0-0.6

0.1-4.5

0.02-0.35

0.5-12

0-1.15

0-0.1

0.2-20

0-0.025

0.7-11

30-31

7-7.6

2.2-6.3

241-1280

0.6-5.3

0-0.05

0-1.2

0.15-3

0.03-0.3

0.5-16

0.03-0.23

0.04-0.4

39-79

0-0.13

24-40

24-27

8-8.3

2.5-8.4

135-713

9.8-11.4

0.01-0.09

0.05-0.5

0.1-0.4

0.3-0.6

0.2-0.5

0.25-0.65

0.00-0.01

7.2-13

0.05-0.18

12-14

24-28

7-8.3

1.1-1.4

13-287

0.4-22.5

0-2.2

0-1.53

0.1-1.3

0.1-0.68

0.3-13

0.07-1.4

0-0.05

0.4-27

0-0.07

5.1-14

22.6-31.6

6.8-8.6

2.7-4.6

65-1230

0.2-6.2

0-0.2

0-0.25

0.01-2

0.01-0.17

1.1-8

0-0.2

0-0.18

12-69

0-0.24

2.9-63

24-28

6.5-7.6

2.3-3.3

64-821

0.2-2.3

0-0.22

0-0.74

0.03-0.58

0.01-0.16

0.1-12

0-0.4

0-0.08

0.1-21

0-0.02

0.5-18

23.5-33

6.6-6.7

1.2-6.1

644-1382

0-0.2

0-0.07

0.1-0.62

0.03-0.94

0.04-1

4-41

0.01-0.18

0.04-0.6

2-21

0.01-0.2

6.0-69

25.5-27.5

7-7.6

4.0-4.1

11-700

0.1-2

0-0.4

0-0.1

0.1-2.2

0-0.28

0.4-19

0-0.86

0-0.1

0.1-26

0-0.01

12-18

30-32

6.7-7.4

2.6-6.5

37-1299

0.0-3.5

0-0.11

0-0.17

0-1.3

0.06-0.9

1.0-19

0.01-0.44

0-0.04

0.2-26

0-0.1

13-67

27

8

6.3

0

0.02

0.37

0.27

1.3

0.9

0.1

0.35

0

2.6

0.04

9.5

24.3.28

7.3-7.5

1.5-2.6

21-31

0.1-0.7

0.04-0.3

0-0.03

0.5-1.22

0.13-0.7

0.7-3.3

0-0.4

0.0-0.07

0.3-1

0-0.03

7.6-14

24.3-29

7.2-8.4

6.0-8.1

1204-5016

13-88

0-0.66

0-0.28

0.013-0.6

0-0.44

0.1-51

0.05-1.44

0.01-0.5

13-68

0.05-0.38

7.3-38

27-31

7.4-7.8

1.1-2.5

1740-3858

4.2-21

0-0.1

0-0.07

0.1-1.7

0.01-0.05

0.0-2.4

0.2-1

0-2.3

157-590

0.16-0.26

11-184

24-25

7

0.2-0.3

11-60

0.1-0.9

0-0.2

0-0.02

0.07-1

0-0.48

0.0-0.6

0-0.5

0-0.12

0.3-6.5

0-0.1

0.4-3.9

27

7.0

0.2-4.7

48-192

0.1-0.4

0.01-0.06

0-0.22

0.05-0.4

0.2-1.16

0-1.3

0.05-1.2

0

0.4-1.3

0-0.02

6.3-56

Data are presented as range of concentrations. All values are in mM, except for COD (mg/L) , temperature (oC), electrical conductivity (mS/cm) and pH. For North-West, South-East (S-E) and South no data for the dry season were obtained

107

around the landfill, except for the southern part. Also some wells in the north, north eastern and eastern part contained lower concentrations, comparable to the reference wells (Table 1). CH4 concentrations (< 1 mM) were consistently higher than in reference wells. CH4 was positively correlated with COD (r = 0.822, p < 0.001), NH4

+ (r = 0.923, p < 0.001), HCO3- (r = 0.949, p < 0.001), and H2S = (r = 0.797, p < 0.001) but negatively with SO4

2- (r = -0.307, p = 0.23). Landfill leachate typically contains high concentrations of salts, which contribute to high electrical conductivity (EC). Therefore, these parameters are often used as indicators for leachate pollution, whereas Cl- is used as a conservative tracer to establish the degree of degradation of organic matter (Van Breukelen et al. 2003). Very high Cl- concentrations (84 mM for the dry season and 38.2 mM for the wet season) were indeed found in landfill leachate, but sometimes also high Cl- and EC values were found in the reference wells (Table 1). For example, in the dry season on one occasion a very high electrical conductivity (46.7 mS/cm) was found for a reference well (Table 1). This high conductivity is indicative for the intrusion of seawater (EC of local sea water 47 mS/cm) (Indonesian Ministry of Environment 2004). High electrical conductivity (EC) and Cl- were found not only in leachate samples but also in many groundwater samples (Table 1). The highest Cl- concentrations in groundwater were found in the N, E and N-E clusters (up to 69 mM). Seawater intrusion should also contribute to high Br- and SO4

2- concentrations (in the absence of significant sulfate reduction). Indeed, Cl- was positively correlated (p < 0.001) with electrical conductivity (r2 = 0.56), SO4

2- (r2 = 0.93) and Br- (r2 = 0.92). These results support the occurrence of seawater intrusion. The maximum SO4

2- concentration in groundwater was found in the NE cluster (41 mM; seawater SO4

2- ~29 mM). Low COD and NH4+ concentrations (< 50 mg/l and < 3.5 mM

respectively) were found for many groundwater samples, such as in the eastern cluster and the reference, suggesting groundwater in these samples was not affected by leachate but seawater. Under the assumption that the higher Cl- concentrations in groundwater samples were indeed due to intrusion of sea water, we calculated the fraction of seawater in groundwater using the formula presented in the Methods section. The highest seawater fraction, 15%, was found for the N-E cluster, while the W region cluster wells contained up to 8% seawater. 3.2. Seasonal variation Hydrochemical parameters for each cluster of observation wells were compared between the dry and wet seasons (Fig. 2). Two-way ANOVA showed significant effects of sampling location (well clusters) for COD (p < 0.01), NH4

+ (p < 0.001), CH4 (p < 0.05) HCO3- (p <

0.001) and PO43- (p < 0.001), but not for other parameters. Seasonal effects were observed for

NH4+ (p <0.05), CH4 (p <0.05), HCO3

- (p < 0.05) and Cl- (p < 0.001). Groundwater from all clusters exhibited higher Cl- concentrations during the dry season (Fig. 2A). SO4

2-, another major component of seawater but not conservative due to the possibility of microbial sulfate reduction, did not reveal a clear seasonal trend (Fig. 2G). The average concentration of Fe2+ in most well clusters showed a seasonal effect (Fig. 2F) similar to Cl-, to a lesser extent it was also observed for Mn2+ (data not shown) . The clearest seasonal effects for indicators of landfill leachate were observed for wells in the landfill. NH4

+ concentration in landfill leachate during the wet season was significantly higher than in the dry season, while NH4

+ concentrations in groundwater wells surrounding

Chapter 5

108

E F G H

0

20

40

60

80

100

120

140 C

l-, m

M

W N N-E E LF REF W N N-E E LF REF Dry season Wet season

abc abc

a ab

a

a

bc

c c c c

c

A

0

1000

2000

3000

4000

CO

D, m

g/L

Dry season W N N-E E LF REF W

N

N N-E E LF REF Wet season

a

a

b bc

d d d

cd cd bc cd bc

B

0

10

20

30

40

50

60

70

NH

4-N

, mM

Dry season W N N-E E LF REF W N N-E E LF REF

Wet season

a

b b b b

b

b b

b

b b b

C

0 0.05

0.1 0.15

0.2 0.25

0.3 0.35

0.4

PO

4, m

M

Dry season W N N-E E LF REF W N N-E E LF REF

Wet season

D

0

100

200

300

400

500

Alk

alin

ity, m

M

Dry season W N N-E E LF REF W N N-E LF E REF

Wet season

a

b b

b b b b b b b b

b 0

0.5

1

1.5

2

2.5

Fe (I

I), m

M

Dry season W N N-E E LF REF W N N-E E LF REF

Wet season

a

ab

bc

ab

a

bc

bc

bc bc bc c

bc

0

5

10

15

20

25

30

SO

4, m

M

Dry season W N N-E E LF REF W N N-E E LF REF

Wet season

bc

bc

a

abc bc

bc

bc bc

bc

abc

ab

c 0

0.5

1

1.5

2

2.5

3

CH

4, m

M

W N E N-E LF REF W N N-E E LF REF Dry season Wet Season

a

b b

b

b b b b b

b b

Hydrochemical characterization of a tropical, coastal aquifer

109

Figure 2. Seasonal and spatial difference in hydrochemical parameters in and surrounding the Keputih landfill, Surabaya. Black bars indicatre landfill leachate samples and grey bars groundwater samples. A. Cl-; B. Chemical Oxygen Demand (COD); C. Ammonia; D. PO4

3-; E. HCO3

-; F. Fe2+; G. SO42-; H. CH4. All data are presented as means, with error bars

indicating the standard deviation. Significant differences (p < 0.05) between season and clusters of wells are indicated by a different character above the bars. For positions of the wells and well clusters, see Figure 1; W, west; N, north; N-E, northeast; E, east; LF, landfill; REF, Reference. the landfill and reference wells were not significant different between seasons (Fig. 2C). The reverse was observed for alkalinity and CH4, these were during the dry season significantly higher in landfill leachate than the wet season, but was not for groundwater taken outside the landfill area (Fig. 2E, H). 3.3. Distribution of redox species The redox status needs to be determined to evaluate natural attenuation (Christensen et al. 1999). The groundwater was anaerobic (dissolved oxygen concentrations were always (well) below 0.6 mg/L). Considerable spatial variation in NO3

- and NO2- concentrations was

recorded in wells in and surrounding the landfill, but concentrations were generally very low, and often below the detection limit. This suggests that aerobic and nitrate reducing processes are not of major importance. Fe2+ concentrations were high (up to 4.5 mM) in the “contaminated” area during the dry season, especially in west, north, north-east and landfill wells but also in the “cleaner” cluster of eastern wells (max. 1.3 mM), in contrast to the reference wells which exhibited a maximum concentration of 0.4 mM. We paid special attention to SO4

2- since it was observed that the concentration of this electron acceptor was abundant, especially due to intruding seawater. Seasonality did not significantly affect the SO4

2- concentration, and only a minor influence of sampling location was observed (Fig. 2G). Average concentrations were up to 15 mM. The high SO4

2- concentrations provide a potentially very high electron accepting capacity. This capacity can be utilized: anaerobic microcosms experiments, using sediment samples collected during well drilling and incubated without or with the addition of glucose, indicated high potentials for iron and especially sulfate reduction (Goeltom et al., unpublished data). H2S, the product of SO4

2- reduction, was detected in leachate (LF cluster) and groundwater from several sites around the landfill, such as N, E, and reference wells. Distribution of sulfide was significantly affected by sampling location (p < 0.05) and season (p < 0.001). Its concentration in the references wells was slightly higher than that found in landfill samples during the dry season whereas the opposite fact was observed during the wet season. In the wet seasons, groundwater from some wells in the northern area (i.e. N1 and N3) exhibited H2S

concentration (max 1.4 mM) comparable to leachate samples. CH4 was observed within the landfill body (LF cluster) at a maximum of 2.3 mM during the dry season but also in “clean” areas like the reference and E wells. Compared to the concentrations in leachate, the CH4 concentration in groundwater from wells around the landfill was lower and at the reference wells it was below the detection limit. In the wet season the highest CH4 concentration (0.12 mM) was detected in the reference well, compared to other well clusters.

Chapter 5

110

4. Discussion 4.1. Hydrochemical characteristics and contamination status The regulatory structure for landfills in the U.S.A. specifies a 30-year post-closure monitoring period during which the landfill is monitored. It is presumed that at the end of this period the landfill and its surrounding environment are stable (Kjeldsen et al. 2002). The guidelines on monitoring are derived from research on landfills in countries with a temperate climate. Temperature and humidity differ quite considerably between temperate climate countries and tropical countries, therewith possibly demanding different monitoring guidelines. In this study, landfill leachate and its infiltration into the subsurface around a closed landfill (4-year post closure) in Surabaya, Indonesia, was monitored because it is important to better understand the subsurface fate of leachate in tropical countries. The composition of Keputih landfill leachate was subject to seasonal variation, containing the highest concentrations of organic matter (COD) and ammonium (ranging from 120 to 5010 mg/L and 247 to 1579 mg/L, respectively) during the wet season. These concentrations exceed the Indonesian national regulatory standard of waste water (for COD: 100 mg/L and ammonium 10 mg/L) (Indonesian Ministry of Environment 1995). High concentrations of COD and ammonium are caused by decomposition of solid organic wastes buried in the past. This also suggests that decomposition of solid waste in the landfill is still occurring. Compared to other tropical, closed landfills (Buyong et al. 2004; Chu et al. 1994; Mangimbulude et al. 2009), the concentration of COD and ammonium in this leachate were higher: e.g. Chu et al. (1994) reported 147 - 1500 mg/L for COD and 65-850 mg/L for ammonium while Buyon et al. ( 2004) (84 - 1637 mg/L for TOC and 1.5 - 1020 mg/L for ammonium). Concentrations were also higher than in the Jatibarang landfill in Semarang, Indonesia, which is still in operation (Mangimbulude et al. 2009). The leachate contaminated the surroundings of the landfill: COD and ammonium concentrations of groundwater in N, NE, E and W areas were higher than in the reference wells. These concentrations also exceeded the national regulatory standards for water quality (25 mg/L for COD and 0.5 mg/L for NH4) (Indonesian National Regulation 2001). This indicates that 4 years after the closure of the Keputih landfill, its leachate has still high potential as source of pollution to the environment. Consequently, landfill leachate should be monitored regularly and it is should be proper handled to protect the groundwater around the landfill. 4.2. The occurrence and impact of seawater intrusion Another reason to investigate the Keputih landfill in Surabaya is that this landfill is close to the coast (< 4 km distance), in a salt marsh, and seawater intrusion could possibly affect the characteristics of leachate and its decomposition. Indeed, we observed the infiltration of seawater. A study conducted by Rachmansyah (Rachmansyah 2001) on municipal waste landfills in Surabaya and its surroundings, also suggested that the area around the Keputih landfill is contaminated by seawater. A major problem in urban areas in most third world countries is finding adequate waste disposal sites. Due to low-income and limitation of available land, coastal areas and salt marshes, which generally have relatively little direct economic value, are often converted to waste disposal sites (Hoornweg et al. 1999). In many third world countries, landfills are been located in coastal areas (Khoury et al. 2000; Olobaniyi and Owoyemi 2006). This is also the

Hydrochemical characterization of a tropical, coastal aquifer

111

case in Indonesia, for example, besides the Keputih landfill, there are at least three other landfills in Surabaya that are located in areas subjected to seawater intrusion (Rachmansyah 2001). Also other major cities in Indonesia such as Padang (Sumatra) and the islands of Bali and Batam have landfills located in coastal areas (Arbain and Sudana 2008; Rofihendra and Trihadiningrum 2010). The occurrence of landfills in coastal areas implies a potential pollution of the marine and coastal environment. Chofqi et al. (2004) reported that since 1999 a coastal aquifer in Marocco has become polluted by an urban landfill. The accumulation of heavy metals such as Hg, Zn and Cu in mangrove sediment in Brazil was influenced by past metal emissions from landfilling surrounding the site (Machado et al. 2002). The groundwater surrounding the Keputih landfill is contaminated by leachate, while also the occurrence of seawater intrusion and seasons affect groundwater quality. Due to fluctuations in seawater intrusion and/or leachate flow over the seasons, the leachate appears to have spread almost all around the landfill. The groundwater at N, NE, E and W regions were more contaminated compared to the groundwater at S and SE regions. The occurrence of leachate at various different positions around the landfill indicates that monitoring the impact of landfill leachate on groundwater will need a large set of wells around landfills affected by seawater intrusion. Salinity affects the occurrence and activity of microorganisms, and may therewith influence landfill leachate decomposition and the spread of leachate contaminants through groundwater. Salinity is a major environmental determinant of microbial community composition (Lozupone and Knight 2007; Herlemann et al. 2011). Lin et al. (2008) observed that increasing salinity depressed the atrazine degrading activity of the microorganisms. Nitisoravut and Klomjek (2005) experimentally investigated and modelled the inhibitory effect of salinity on biological oxygen demand (BOD) removal. Salinity inhibited the metabolism of microorganisms in a wetland environment subjected to leachate, which might be critical for the proper functioning and maintenance of the system. On the other hand, also in marine environments there is generally a high potential for degradation of contaminants (e.g. Head et al. 2006). 4.3. Potential for the natural attenuation of landfill leachate Landfill leachate contains high concentrations of humic acids. These organics are not easily biodegraded (Calace and Petronio 1997) and may enhance the solubility of hydrophobic compounds through complexation. Humic acids could modify bioavailability and biotoxicity of hazardous compounds, thereby impacting biodegradation or bio-uptake as well (Fan et al. 2007). The fate of organic pollutants in contaminant plumes depends on the presence of redox components such as dissolved oxygen, NO3

-, Fe(III)-oxides and SO42- (Baun et al. 2003).

Thus, evaluation of natural attenuation can be accomplished by determining the occurrence and loss of electron acceptors (Baun et al. 2003; EPA 1999). Seawater intrusion into fresh groundwater aquifers on the one hand affects coastal groundwater quality. On the other hand, seawater intrusion provides a large source of redox species such as sulfate to support microbial activity, which is relevant if groundwater is polluted with landfill leachate. This study showed that high SO4

2- concentrations were observed in landfill leachate and groundwater, in comparision to other redox elements. Jørgensen (1982) reported that the microbial decomposition of organic matter coupled with the reduction of sulfate is an important mechanism governing carbon and energy in many

Chapter 5

112

anaerobic environments. Cozzarelli et al., (2000) observed the anaerobic degradation of landfill leachate in an aquifer in the U.S.A. under sulfate reducing conditions. Tian et al. (2005) and Wu et al., (2009) reported that sulfate reducing bacteria played important roles in leachate-polluted aquifers along the shore of the East China Sea. In our study, high concentrations of H2S and Fe2+ were found in landfill leachate and groundwater, indicating that sulfate-reduction and iron-reduction were predominant processes in groundwater. Stoichiometrically, 1 mM of COD needs 0.5 mM of SO4

2- for its complete oxidation. Based on the measured COD and SO4

2- levels, the availability of sulfate in groundwater and leachate will not be enough to reduce the COD concentration to an acceptable level (< 100 mg/L), despite seawater intrusion. However, this outcome also depends on the mixing proportions of saline groundwater and leachate. If leachate spreads further, it will mix with a larger volume of saline and sulfate containing groundwater, eventually possibly leading to acceptable COD reductions. Moreover, biodegradation of organic matter occurred not only by sulfate reducing bacteria, but could also occur through other anaerobic processes such as methanogenesis. Laboratory microcosm experiments, with Keputih-contaminated sediments, indicated a high intrinsic potential for anaerobic redox process for Keputih sediments, especially sulfate and iron reduction (Goeltom et al., unpublished results). Overall, there is a need to further study the impact of landfill leachate and salinity on the occurrence and behavior of subsurface microbial communities in tropical countries, as in many tropical countries coastal areas are impacted by leachate mixing with infiltrating seawater. For example, molecular approaches based on targeting functional genes such as dissimilatory sulfate reduction (apsA, dsrAB, aprBA) genes can been used to investigate the diversity of anaerobic sulfate (sulfite) reducers (Cook et al. 2008; Zhang et al. 2008). Acknowledgments The authors would like to thank WOTRO Science for Global Development and the V. Rutgers fund of the VU University Amsterdam, The Netherlands, for funding the work described here. . References Appelo, C.A.J., Postma D., 2005. Geochemistry, groundwater and pollution. Balkema,

Rotterdam, the Netherlands. Arbain, N.K.M., Sudana, I.B., 2008. The influence of the Suwung landfill leachate on shallow

ground water quality in the village Pedungan surrounding the city of Denpasar (in Indonesian). Echotropic 3, 55-60.

Baun, A., Rietzel, L.A., Ledin, A., Christensen, T.H., Bjerg, P.L., 2003. Natural attenuation of xenobiotic organic compounds in a landfill leachate plume (Vejen, Denmark). Journal of Contaminant Hydrology, 65, 269-291.

Buyong, F.H., Abdul Kadir, M.S., Darus, F.M., 2004. Comparison of selected parameters of leachate in different age of closed landfill. In: e-Proceeding of 17th. Analysis Chemistry Malaysia Symposium, Swiss-Garden Resort & Spa, Kuantan, Pahang, 24-26 August 2004. pp 699-708

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Calace, N., Petronio, B.M., 1997. Characterization of high molecular weight organic compounds in landfill leachate: Humic substances. Journal of Environmental Science and Health, part A 32, 2222-2244.

Chaillan, F., Le Fleche, A., Bury, E., Phantavong, Y., Grimont, P., Saliot, A., Oudot, J. 2004. Identification and biodegradation potential for tropical aerobic hydrocarbon-degrading microorganisms. Research in Microbiology 155, 587-595.

Chofqi, A., Younsi, A., Lhada, E.K., Mania, J., Murdy, J., Veron, A., 2004. Environmental impact of an urban landfill on coastal aquifer (El-Jadida, Marocco). Journal of African Earth Science 39, 509-516.

Christensen, T.H., Kjeldsen, P., Albrechtsen, H.J., Heron, G., Nielsen, P.H., Bjerg, P.L., Holm, P.E. 1999. Attenuation of landfill leachate pollutants in aquifers. Critical Reviews in Environmental Science and Technology 24, 119-202.

Christensen, T.H., Kjeldsen, P., Bjerg, P.L., Jensen, D.L., Christensen, J.B., Baun, A., Albrechsten, H.-J,, Heron, G., 2001. Biogeochemistry of landfill leachate plumes. Applied Geochemistry 16, 659-718.

Chu, M.L., Cheung, K.C., Wong, M.H., 1994. Variation in the chemical properties of landfill leachate. Environmental Management 18, 105-117.

Cline, J.D., 1969. Spectrophotometric determination of hydrogen sulfide in natural waters. Limnology Oceanography 14, 454-458

Cook, K.L., Whitehead, T.R., Spence, C., Cotta, M.A., 2008. Evaluation of the sulfate-reducing bacterial population associated with store swine slurry. Anaerobe 14, 172-180.

Cozzarelli, I., Sulfita, J., Ulrich, G., Harris, S., School, M., Schlottmann, J., Christenson, S., 2000. Geochemical and microbiological methods for evaluating anaerobic processes in an aquifer contaminated by landfill leachate. Environmental Science and Technology 34, 4025-4033.

Cozzarelli, I.M., Suflita, J., Ulrich, G., Harris, S., Scholl, M.A., Schlottmann, J.L., Jaeschke, J.B., 1999. Biogeochemical processes in a contaminant plume downgradient from a landfill, Norman, Oklahoma. In: Morganwalp DW, and Buxton, H.T., eds. U.S. Geological Survey Toxic Substances Hydrology Program, Charleston, South Carolina, 1999, pp 521-530.

Endah, N., Pudjiastuti, L., 1995. The influence of waste from the city of Surabaya at the Sukolilo Landfill on water quality of resident wells surrounding the landfill (in Indonesian). MSc thesis, Surabaya Technology Institute.

EPA, 1999. Use of monitored natural attenuation at superfund, RCRA corrective action and underground storage tank sites. Washington D C.

Fan, H.J, Chiu, T., Yang, H.-S., Chen, W.-C., Furuya, E., 2007. Removal of humic acid, fulvic acid and non-humic substances. Journal of Environmental Engineering and Management 17, 325-331.

Head, I.M., Jones, D.M., Röling, W.F.M., 2006. Marine microorganisms make a meal of oil. Nature Reviews Microbiology 4, 173-182.

Herlemann, D.P.R., Labrenz, M., Jürgens, K., Bertilsson, S., Waniek, J.J., Anders, F., Andersson, A.F., 2011. Transitions in bacterial communities along the 2000 km salinity gradient of the Baltic Sea. ISME Journal 5, 1571-1579.

Hoornweg, D., Thomas, L., Varma, K., 1999. What a waste: solid waste management in Asia. Urban Development Sector Unit East Asia and Pacific Region. The World Bank., Washington DC.

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Indonesian Ministry of Environment, 1995. Standards and control on the quality of industrial waste water (in Indonesian). Jakarta, Indonesia.

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6 General discussion

The previous chapters have presented new data on the biogeochemistry of groundwater and leachates associated with two landfills in Indonesia: Jatibarang (close to Semarang, Central Java), and Keputih (close to Surabaya, East Java), as well as on the composition of microbial communities contributing to natural attenuation of polluted leachates. The present chapter will summarize and discuss these data by addressing two issues: (i) leachate composition and how to control it, and (ii) the potential for natural attenuation processes in groundwater. Leachates from tropical landfill Leachate from municipal landfill usually contains a large number of contaminants with concentrations varying from time to time and from site to site due to factors such as waste composition, type of waste, age of the landfill, weather conditions, seasonal influences, and landfill operation (Chu et al., 1994; Lema et al., 1998; Fadel-El, 1998; Christensen et al., 2001). Leachates from tropical landfills are characterized by a high organic matter content (measured as chemical oxygen demand, COD) and high ammonium concentrations (Chu et al., 1994; Aluko et al., 2003, Fan et., 2005, Chapter 2). In the case of leachate from the Jatibarang and the Keputih landfills, COD and ammonium concentrations were found to be higher than the Indonesian national standard for waste water (Chapter 2 and Chapter 5). Decomposition of buried materials in a landfill will always release organic matter into the leachate. Kjeldsen et al. (2002) reported that waste decomposition was still ongoing 30 years after closure of the dump. This implies that long-term monitoring of a landfill is important for management of effects on the environment. In addition to organic matter and ammonium, leachates may also contain high concentrations of sulfate, iron (II), calcium, hydrogen carbonate and chloride. In addition, low concentrations of dissolved oxygen, nitrite and nitrate are usually observed. These inorganic components are all associated with the biodegradation processes in a landfill. Berge et al. (2005) reported that landfill leachates were rich in organic matter and low in nitrate and nitrite, but ammonia remained an issue. In landfills monitored for a longer time, usually no decreasing trend with time is observed for ammonium. Ammonium was considered the most significant hazardous component of landfill leachate on the long term (Robinson, 1995; Krumpelbeck and Ehrig, 1999; Christensen et al., 1999). Microbial degradation of proteins and other nitrogenous compounds contributes to the release of ammonium in the leachate. Data from a laboratory experiment reported in this thesis showed that ammonification rates in leachate were higher than the rates of nitrification and anaerobic ammonia oxidation. These observations could be the reason why ammonium remains to be present in the leachate at a high level (Chapter 3).

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The above facts indicate that the Jatibarang and Keputih landfills have a high potential to pollute groundwater and surface water. The risk of groundwater pollution is probably the most severe environmental impact from landfills (Kjeldsen et al., 2002). In the case of the Keputih landfill, it is clear that not only groundwater under the landfill, but also groundwater at sites surrounding the landfill was contaminated with leachate components (Chapter 5). Surface water pollution caused by leachate from landfills has also been observed, although relatively few cases have been described in the literature (Kjeldsen et al., 2002). In the case of Jatibarang, there was not a negative impact of leachate on river Kreo water quality due to significant dilution (Chapter 2). However, long-term accumulation of leachate contaminants along the river water may affect aquatic life. A critical point of tropical leachate hydrochemistry are seasonal influences. The amount of water infiltrating in the rainy season (Tränkler et al., 2005) increases the volume of leachate produced (Chapter 2). In other words, the volume of leachate is highly dependent on the rainfall that falls on a landfill if it is not covered. Consequently, it is important to understand the effect of leachate generation and its characteristics as a function of local climatic variation, landfill design and its operation (Visvanathan et al., 2002; Fan et al., 2005). Leachate management in practice Landfill leachate management is part of municipal solid waste management that can be used to control leachate generation, collecting and treatment, in order to reduce the potential risks for the environment (Damiecki et al., 2007; Johannessen and Boyer, 1999). In the last few decades, landfill leachate management has become an important issue of concern for many countries in the world (Cointreau, 1982; Doan, 1998; Agunwamba, 1998). In developing countries like Indonesia, with increasing population growth and consumption rates, landfill leachate management has also become a major problem (Johannessen and Boyer, 1999; Liermann, 2009; Meidiana and Gamse, 2010). One of the issues in waste management in Indonesia is lack of finance. The local government has policies on fee retribution and there is a local governmental budget for municipal solid waste management. However, the amount of the whole contribution is still little (about 2% of the total local budget) and does not cover the expenses needed for proper waste management; the collection rate of the retribution amounts only 40 - 50 % of the revenue. Such a limited allocation to the waste sector leads to a low level of service by the municipal solid waste management (Meidiana and Gamse, 2010). Also in the case of the Jatibarang landfill, leachate generation, collecting and treatment are inadequately managed due to financial limitations. For example the system of channels from the landfill to the leachate collection ponds causes leachate to flow out of the pond. The leachate ponds were not well constructed, and sometimes leakage occurrs through the bottom or through the walls (Chapter 2). The leachate treatment system does not run according to operating system design criteria (Zaman et al., 2012). Conventional methods for contaminated groundwater treatment involve pumping groundwater to the surface, followed by treatment and disposal (Blowes et al., 1997). However, Mackay and Cherry (1989) have shown that these methods are expensive and in many cases, are ineffective in achieving the proposed level of clean-up. Among the environmental remediation methods, natural attenuation is receiving attention as an option for removing organic contaminants from an aquifer (Bekins et al., 2001). The term natural attenuation refers to the intrinsic bioremediation capacity of the groundwater

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ecosystem (Röling and Van Verseveld, 2002; Scow and Hicks, 2005). Microorganisms are the principal mediators of natural attenuation of many pollutants, such as organics, metals and inorganic nitrogen compounds (Christensen et al., 2001). In general microbial degradation of organic contaminants can occur naturally if suitable electron donors, electron acceptors and nutrients are available (Röling and Van Verseveld, 2002; Scow and Hicks, 2005). Leachate microbial communities at the Jatibarang landfill In general, leachate contains a number of essential chemical components including organic carbon, nutrients (nitrate, carbonate and phosphate), electron acceptors (dissolved oxygen, nitrate, manganese (IV), iron (III), sulfate and other compounds). These components may form a special niche for certain micro-organisms. In other words, leachate is like a natural medium for many micro-organisms. A number of studies have reported on the variable composition of microbial communities in landfills (i.e. Archaea, Bacteria, and Fungi) (Ludvigsen et al., 1999; Röling et al., 2001; Boothe et al., 2001; Christensen et al., 2001; Brad et al., 2008). In the case of leachate from Jatibarang, the microbial communities were analyzed directly using molecular approaches such as denaturating gradient gel electrophoresis of PCR-amplified taxonomic marker genes (DGGE) to determine the diversity of the community, and by targeting functional genes. In addition, through laboratory experiments microbial activities contributing to natural attenuation of landfill leachate were assessed. Chapter 4 showed a high diversity of microbial communities in the leachate which varied over time. One of the dominant microbial groups found in the Jatibarang leachates were anaerobic aromatics-degrading bacteria. The availability of electron acceptors (sulfate) was high enough to sustain the ongoing activities of anaerobic communities. Often, electron acceptor availability is considered a dominant factor controlling the activity of organisms (Harris et al., 2005). Cozzarelli (2001) reported that sustainability of natural attenuation in contaminated aquifers depends on the availability of electron acceptors. From the laboratory experiments, it is clear that heterotrophic and autotrophic microbes are responsible for the transformation of nitrogen, e.g. denitrification, nitrification and anaerobic ammonia oxidation (Chapter 4). Five pure isolates were able to grow on medium containing nitrite at 2000 mg N L-1 (Chapter 4). Profiles of microbial communities and the activities of nitrogen transformations in the Jatibarang landfill leachate gave some insights in the role of microbial communities in natural attenuation of nitrogen compounds. However, for other compounds in the leachate, e.g. aromatics such as benzene, toluene, ethylbenzene and xylene (jointly known as BTEX) little is known about the role of degrader communities. Another important question is the extent to which removal of organics in run-off water varies from one place to another. Finally, the role of biofilms may be underrated in the present research. At Jatibarang, biofilms were seen at the base of the channel through which leachate flowed to the treatment ponds. The question is how do communities of microbes organized in biofilms contribute to the nitrogen transformations in Jatibarang leachate? These questions will be considered in further investigations. To improve the Jatibarang landfill management several points should be considered: (a) Leachate generated from the landfill should be controlled to prevent excessive leachate production during the wet season. An appropriate top cover design is one of the main

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parameters to consider for control of leachate generation and maintenance of a water balance under seasonal variation (Tränkler et al., 2005). (b) The leachate canal system should be improved and adequate collecting ponds with respect to minimum and maximum leachate volume during the dry and wet seasons, should be designed (Visvanathan et al., 2002; Asian Regional Research Programme on Environmental Technology, 2004; Chapter 2). (c) The existing leachate treatment plant does not remove organic and ammonium to an acceptable level, therefore, improvements in the leachate treatment plant and the operational system need to be considered. This can be achieved by increasing the residence time of leachate in the leachate collection system and by increasing its volumetric capacity (Chapter 2). To remove organic matter and ammonium, the leachate treatment should proceed through two stages, aerobic and anaerobic. In the first stage, introducing oxygen to a level above 2 mg L-1 is needed, in order to increasing nitrification and organic biodegradation processes. In the second stage nitrate is converted to nitrogen gas through denitrification or anammox processes (Chapter 2 and 3; Wang et al., 2007). Other alternatives include a combination of stable high ammonium removal using a nitrite reactor system (SHARON) and anammox (Schmidt et al., 2003; Paredes et al., 2007). Potential for natural attenuation of groundwater surrounding the Keputih landfill: options for remediation Evaluation of the hydrochemistry of the coastal aquifer underlying the Keputih landfill revealed that the groundwater is affected not only by landfill leachate but also by seawater intrusion (Chapter 5). These results illustrate that Surabaya City is facing a serious problem with groundwater contamination by salt water, as described by previous researchers (Suryono et al., 2008; Kartina, 2011). This may affect drinking water resources, as there is no appropriate remedial option to remove salt from groundwater. In the case of contaminated groundwater around the Keputih landfill, the presence of redox elements such as nitrate, sulfate and intermediary metabolites such as methane, sulfide, nitrite, iron(II) and manganese(II) were evidence for the occurrence of natural attenuation (Chapter 5). Due to oxygen depletion (< 0.6 mg L-1), it was clear that biodegradation of organic contaminants in the groundwater occurred under anaerobic conditions. In this case, sulfate is the dominantly redox element because of its high availability in groundwater due to a long-term tidal effect (Chapter 5). Also a high concentration of iron(II) was observed in groundwater. This indicates that sulfate reducing bacteria and iron reducing bacteria could play an important role in the degradation of organic contaminants in the groundwater at Keputih. Studies on a contaminated aquifer near the Banisveld landfill (The Netherlands) showed that Geobacteraceae (a family of Deltaproteobacteria) are a dominant group of bacteria with sulfur reducing and iron reducing capacities, that play an important role in organic degradation (Röling et al., 2001; Van Breukelen et al., 2004; Lin et al., 2005). A study of the leachate plume from the Norman landfill in Oklahoma revealed that contaminant organic matter degradation was mainly conducted by sulfate-reducing bacteria (Cozzarelli, 1999). Many studies on microbial composition and their dynamics in groundwater contaminated by landfill leachate have been done, however due to the differences in ecological conditions every aquifer shows different microbial dynamics (Pickup et al., 2002; Johnson et al., 2003;

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Tian et al, 2005). The question arises whether the sulfate reducers are the dominant community in contaminated groundwater surrounding the Keputih landfill? Until now little is known about microbial community structures and their dynamics in aquifers under a landfill affected by seawater intrusion. To obtain a comprehensive view on the possibilities of natural attenuation integrated studies are needed of aquifer microbial communities in relation to environmental conditions and geochemical processes. Conclusions The management of landfills in tropical countries remains a great challenge. Management should be aimed at preventing contamination of groundwater, streams and rivers. The challenge is complicated by the enormous seasonal variation in rainfall, causing widely varying surface run-offs. My investigations at two sites in Java, Indonesia have shown that microbial communities contributing to degradation are in place, and their activity can be stimulated by appropriate options in landfill design and operation, e.g. constructing leachate treatment ponds with alternating anaerobic and aerobic conditions. Special attention should be paid to the often very high ammonium concentrations and aromatics such as phenolic compounds. The clever use of microbial activity for natural attenuation is a good strategy since more sophisticated waste treatment processes are usually not available in tropical countries due to financial limitations. It is recommended that natural attenuation be accompanied by appropriate long-term monitoring to make sure that contamination of surface water and groundwater is prevented as much as possible. References Agunwamba, J., 1998. Solid waste management in Nigeria: problems and issues.

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Hydrochemie en natuurlijke afbraak in lekwater uit vuilstortplaatsen in de tropen

Samenvatting Het storten van huisvuil wordt nog steeds op grote schaal toegepast in de tropen, vaak zonder speciale maatregelen om vervuiling van het milieu te voorkomen. De hoeveelheid lekwater die uit een vuilstort stroomt, en de chemische samenstelling ervan, vertonen een grote variatie die voor een groot deel bepaald wordt door de geometrie van de belt, de ouderdom van het afval, de neerslag, de luchtvochtigheid en de temperatuur. In de tropen variëren de weersomstandigheden sterk met de seizoenen; hiermee moet rekening gehouden worden bij de controle op lekwater. Een combinatie van fysische, chemische en microbiële processen kan bijdragen aan verontreinigende afvalstromen vanuit een vuilstort; een beter begrip van deze processen is nodig voor een goed beheer van de vuilstort. In dit proefschrift wordt in het bijzonder gekeken naar (a) de kenmerken van lekwater, de veranderingen in samenstelling over de seizoenen en de invloed op het oppervlaktewater en het grondwater, (b) de door micro-organismen uitgevoerde transformaties van stikstof en de afbraak van fenolen in het lekwater, en (c) seizoensveranderingen in de microbiële gemeenschappen en beoordeling van hun vermogen tot natuurlijke afbraak. Hoofdstuk 1 geeft een algemene inleiding in de problematiek van afvalbeheer, met bijzondere aandacht voor de praktische problematiek van huishoudelijk afval in Indonesië en daarmee verbonden sociale aspecten. Ook worden de algemene kenmerken van lekwater uit vuilstorten beschreven en de hydrochemische parameters die vaak gebruikt worden om percolaten te karakteriseren en te beoordelen. Microbiële aspecten van de vorming van lekwater en de afbraak van vervuilende stoffen worden geïntroduceerd, vooral met betrekking tot de stikstofcyclus en de biodegradatie van fenolische verbindingen. Ook worden de onderzoekslocaties, de onderzoeksvragen en de gebruikte methodes in dit hoofdstuk gepresenteerd. In hoofdstuk 2 wordt het bovengronds lekwater van de vuilstort Jatibarang, nabij Semarang bestudeerd en de invloed daarvan op een nabije rivier. Het ging daarbij vooral om de seizoensvariatie in de samenstelling van het lekwater en het optreden van natuurlijke afbraak. De mogelijkheden voor natuurlijke afbraak en de bijdrage van biodegradatie langs het stroompad naar de rivier werden maandelijks gemeten op basis van een reeks lekwaterparameters zoals het gehalte organische stof en redox-gerelateerde componenten. Elektrische geleidbaarheid en de concentraties van BOD, COD, opgelost organisch stikstof, ammonium, sulfaat en calcium waren gedurende het droge seizoen hoger (tot een factor 2,3) dan in het regenseizoen. De verblijftijd van afvalwater in het systeem van basins waardoor het lekwater stroomde, was minder dan 70 dagen. Veldmetingen lieten zien dat er gedurende deze periode amper enige afbraak van organische stof en ammoniumverwijdering plaatsvond (minder dan 25%). Echter, de potentie voor biodegradatie van organische stof en ammoniumverwijdering werd duidelijk aangetoond in proeven gedurende 700 dagen in het laboratorium (meer dan 65%

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verwijdering). Er werd geen significante invloed van lekwater op de kwaliteit van het rivierwater geconstateerd, maar de aanwezigheid van zware metalen en coliforme bacteriën moet op de lange termijn in de gaten gehouden worden. Verdunning was het belangrijkste proces dat verdwijnen van verontreiniging bepaalde, maar dit is geen optimaal afbraakmechanisme omdat verdunning niet de absolute hoeveelheid van verontreiniging in de rivier vermindert. Ammonium is een van de belangrijkste toxicanten in lekwater en een kritische component op de lange termijn. Lekwater van de Jatibarang vuilstort bevat ammonium in concentraties variërend van 376 tot 929 mg stikstof per liter. Het grote verschil in de stikstofflux tussen het natte en droge seizoen bemoeilijkt de controle op de biologische stikstofverwijdering uit het lekwater van de vuilstort. Een beter begrip van de seizoensvariatie in de potentie voor stikstofmineralisatie, nitrificatie, denitrificatie en anammox in het lekwater zou kunnen helpen om het beheer van de vuilstort te verbeteren en te voorkomen dat het oppervlaktewater verontreinigd wordt met ongecontroleerde lozingen van lekwater met hoge concentraties van het toxische ammonium. Daarom werden in hoofdstuk 3 de potenties voor stikstofomzettingen bepaald. Monsters genomen uit hetsysteem van basins voor opvang van bovengronds lekwater gedurende een aantal seizoenen werden gebruikt om synthetische media te inoculeren in het laboratorium. De aerobe capaciteit voor ammoniumoxidatie (< 0.06 mg N L-1 h-1) was honderd keer lager dan die van de ammonificatie en anaerobe stikstofomzettingsprocessen, die onderling vergelijkbaar waren. Effecten van het seizoen werden alleen waargenomen voor aerobe nitrificatie en anammox, en deze effecten waren relatief gering. Over het geheel genomen was de potentie voor volledige nitrificatie in het lekwater van Jatibarang het laagst van alle stikstoftransformatieprocessen. Om de kwaliteit van het lekwater te verbeteren en een aanvaardbaar niveau van anorganisch stikstof te bereiken voor lozing op de rivier de Kreo, moet een tweestaps aeroob/anaeroob systeem bekeken worden, waarin beter gebruik gemaakt wordt van het huidige systeem voor opvang van bovengronds lekwater. Biologische stikstofverwijdering via nitrificatie van ammonium, gevolgd door denitrificatie van nitraat wordt vaak geadviseerd voor de behandeling van lekwater uit een vuilnisbelt met hoge ammoniumconcentraties, en dit is ook van toepassing op Jatibarang. Natuurlijke afbraakprocessen zijn afhankelijk van de gezamenlijke potenties van micro-organismen om verontreinigingen af te breken onder de gegeven omstandigheden. Biodegradatie is vaak het primaire mechanisme voor verwijdering van contaminanten uit een verontreinigd watervoerend pakket in de ondergrond. Daarom stelde ik me in hoofdstuk 4 ten doel om meer inzicht te verwerven in de dynamiek en de diversiteit van microbiële gemeenschappen betrokken bij natuurlijke afbraak, en hun mogelijkheden om giftige organische verbindingen, zoals fenolen, af te breken. Met deze kennis zou het mogelijk moeten worden om de microbiële processen beter te voorspellen en hun bijdrage te verhogen. De temporele variatie in de microbiële gemeenschappen van Jatibarang werd bestudeerd door profielen te maken op basis van gelelectroforese met een denaturerende (DGGE) en deze in verband te brengen met de hydrochemie. De aanwezigheid van anaerobe afbrekers van aromaten werd vastgesteld met een kweek-onafhankelijke PCR-benadering gericht op het aantonen van het bamA gen, dat codeert voor een enzym betrokken bij de anaerobe benzoyl-CoA-afbraakroute. Ook werd de feitelijke potentie voor aerobe en anaerobe fenolafbraak bepaald in cultures. Voor bijna alle monstertijdstippen tussen het droge seizoen van 2004 en het natte seizoen van 2006 werd een hoge diversiteit waargenomen met in het algemeen meer dan 20 bandjes in de DGGE-profielen. De gemeenschapsprofielen varieerden tussen de

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tijdstippen, maar nauwkeurige bestudering leerde dat ze niet compleet veranderden en dat de meeste bandjes aanwezig bleven op elke tijdstip. Echter tussen het natte seizoen van 2006 en het droge seizoen van 2007 was er een plotselinge terugval in microbiële diversiteit. Langs het stroompad van de belt naar de rivier werden verschuivingen in de microbiële gemeenschap waargenomen. Monsterplaatsen bij het begin van het stroompad lieten vaak een grote gelijkenis zien in de profielen, onafhankelijk van het jaar of het seizoen. De gemeenschappen halverwege het stroompad leken nog behoorlijk veel op die aan het begin, maar hadden een aantal extra bandjes. De profielen dicht bij de rivier en in de rivier waren in het algemeen meer verschillend van degene eerder in het stroompad, ondanks het feit dat ze een aantal bandjes daarmee gemeenschappelijk hadden. BamA genen werden op allerlei tijdstippen en plaatsen gevonden, maar restrictieanalyse liet zien dat de diversiteit van aromatenafbrekers laag was. Met kweekexperimenten in het laboratorium werd aangetoond dat micro-organismen in staat waren om fenol te verwijderen met een efficiëntie tot 70%, binnen 60 dagen incubatie onder sulfaat- en nitraatreducerende condities. Veldgegevens lieten zien dat afbraak van fenol en organische fenolische verbindingen mogelijk plaatsvindt onder sulfaatreducerende condities. In hoofdstuk 5 werd een andere vuilstort bestudeerd, de Keputih-vuilstort in Surabaya, dicht bij de kust van Java. Deze vuilstort wordt niet meer gebruikt sinds 2001. De aandacht ging vooral uit naar de kwaliteit van het grondwater onder de belt. Veldmetingen lieten zien dat de elektrische geleidbaarheid en de COD hoog waren, met hoge concentraties van ammonium, HCO3

-, SO42-, Fe2+ en Cl-. De meeste componenten werden ook hier significant

beïnvloed door het seizoen, behalve COD en SO42-. Het grondwater in bijna alle

grondwaterpeilbuizen rondom de vuilstort vertoonde tekenen van menging met percolaat. Op basis van de analyseresultaten kon ook geconcludeerd worden dat het grondwater onder de belt beïnvloed wordt door binnendringing van zeewater. De hoge abundantie van sulfaat in het grondwater betekent een grote potentie voor biodegradatie van organische verbindingen. Hoofdstuk 6 geeft een algemene discussie en recapitulatie van de conclusies van het proefschrift. In zijn algemeenheid kan gesteld worden dat de concentraties van ammonium en organische verontreinigingen in percolaten van de vuilstortplaatsen hoger zijn dan de Indonesische normen voor afvalwater. De hydrochemie van het lekwater in Jatibarang en Keputih staat sterk onder invloed van de seizoenen. Seizoenseffecten zijn een kritische factor bij de ontwikkelingen van een het beheersysteem voor de vuilstort. Het beheer moet gericht zijn op het voorkomen van verontreiniging van grondwater, beken en rivieren. De uitdaging wordt bemoeilijkt door de enorme seizoensvariatie in de regenval die een sterk variërende afstroom via het oppervlakte veroorzaakt. Er kan geconcludeerd worden dat de potentie voor natuurlijke afbraak zeker aanwezig is in beide vuilstortplaatsen, ondanks de grote afvalstromen. Deze potentie moet optimaal benut worden, waarbij met het seizoen variërende factoren meegenomen moeten worden. Wetenschappelijke kennis over de door micro-organismen gekatalyseerde omzettingen en afbraakroutes kan helpen bij het ontwikkelen van optimale strategieën om verontreiniging van grondwater en oppervlaktewater te voorkomen.

129

Hydrochemistry and natural attenuation of leachate from tropical landfills

Summary Landfilling, as a system to discard municipal waste, is still widely used in the tropics, often without appropriate protection against pollution of its surroundings. Leachate generation from landfills, and its chemical composition, varies widely and is significantly influenced by landfill geometry, the type and age of waste, seasonal precipitation, moisture and temperature. In the tropics, climate conditions vary strongly with seasons and need to be taken into account with regard to leachate management. A combination of physical, chemical and microbial processes can contribute to the attenuation of pollution by landfills, but further understanding of these processes is essential in order to allow for better management of landfills. This study focuses on several aspects: (a) the characteristics of landfill leachates, their seasonal changes and the impact on surface and groundwater pollution around the landfill, (b) the microbe-mediated nitrogen transformations and phenol degradation in landfill leachate and (c) seasonal changes in microbial communities in landfill leachate and assessment of their contribution to natural attenuation. Chapter 1 provides a general introduction on solid waste management, with special emphasis on solid waste management practices in Indonesia, and its association with social aspects. Also, it describes the general characteristics of leachate, and the hydrochemistry parameters that are often used for characterization and evaluation. Microbiological aspects of leachate generation and attenuation, particularly nitrogen transformation and biodegradation of phenolic compounds, are introduced. The location of research sites investigated in this study, research questions and methodological strategies are presented in this chapter. In Chapter 2, the surface landfill run-off from the Jatibarang landfill, near Semarang, and its impact on a nearby river has been studied to understand seasonal landfill leachate characteristics and to determine the occurrence of natural attenuation. The potential for attenuation and the contribution by biodegradation along the flow path to the river was determined by measuring monthly a range of leachate parameters, such as organic matter content and redox-related components. The results showed that the composition of leachate was greatly influenced by season. Electrical conductivity and concentrations of BOD, COD, organic nitrogen, ammonia, sulfate and calcium during the dry season were higher (up to 2.3-fold) than in the wet season. The residence time of waste-water in a leachate collection system at the site was less than 70 days. Field measurements showed that during this period hardly any biodegradation of organic matter and ammonia occurred (less than 25%). However, the potential for biodegradation of organic matter and ammonia removal was clearly revealed during 700 days of incubation of leachate in the laboratory (over 65%). No significant impact of landfill leachate on river water quality was observed; however, the presence of heavy metals and coliform bacteria need to be considered over the long-term. Dilution was the major natural attenuation process acting on leachate. However, it is not an optimal attenuation mechanism because dilution does not decrease the absolute load of pollutants discharged into the river. Ammonium is one of the major toxic compounds and a critical long-term pollutant in landfill leachate. Leachate from the Jatibarang landfill contains ammonium in concentrations ranging from

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376 to 929 mg N L-1. The large differences in nitrogen flux between wet and dry seasons provides a difficulty in controlling biological nitrogen removal in landfill leachate. A better understanding of seasonal variation on the potential for nitrogen mineralization, nitrification, denitrification, and anammox in leachate could help to improve the management of tropical landfills, and prevent surface water quality deterioration due to uncontrolled discharge of leachate containing high concentrations of toxic ammonium. Therefore, nitrogen transformation potentials were determined in Chapter 3. Seasonal samples from leachate collection treatment ponds at Jatibarang were used as inocula to feed synthetic media in order to determine potential rates of nitrogen transformations in the laboratory. Aerobic ammonium oxidation potential (<0.06 mg N L-1 h-1) was more than a hundred times lower than ammonification and anaerobic nitrogen transformation processes, which were of the same order of magnitude. Effects of season were only observed for aerobic nitrification and anammox, and were relatively minor: rates were up to three times higher in the dry season. Overall, the potential of full nitrification in the Jatibarang landfill leachate was the lowest of all nitrogen transformation processes. In order to improve the quality of the leachate, and to enable an acceptable concentration of inorganic nitrogen for the discharge of leachate into the river Kreo, implementing a two-step aerobic/anaerobic system needs to be considered to make more optimal use of the current leachate collection system. Biological nitrogen removal through a two-step process of nitrification and denitrification is usually suggested for the treatment of landfill leachate with high ammonium concentrations. Natural attenuation processes rely on the collective abilities of microorganisms to degrade pollutants under prevailing environmental conditions. Biodegradation is a most often the primary mechanisms for contaminant destruction in the polluted aquifer. Therefore, Chapter 4 aimed to gain insight into the dynamics and diversity microbial communities present in landfill leachate and their potentials to degrade toxic organic compounds, like phenol. This knowledge would aid in understanding, predicting and enhancing microbial processes involved in natural attenuation. Temporal variation in microbial communities over seasons at Jatibarang were studied using denaturing gradient gel electrophoresis (DGGE) profiling, in relation to leachate hydrochemistry. The presence of anaerobic aromatic degraders was determined by a cultivation-independent PCR approach in which the presence of a gene (bamA) from the anaerobic benzoyl-coA degradation pathway was determined, and the actual potential for aerobic and anaerobic phenol degradation by a culturing approach were measured. Between the dry season in 2004 and the wet season in 2006 a high diversity was observed for nearly all sampling points, with in general over 20 bands visible in the DGGE profiles. Community profiles differed between time points, although close inspection of the DGGE profiles indicated that community profiles at a particular sampling location did not change completely over time, some bands remained present. However, between the wet season in 2006 and dry season in 2007 diversity suddenly dropped strongly. Along the flow path from landfill to river shifts in community structure were observed. Sampling locations at the start of the surface flow path from landfill to river often showed high similarity in community profiles, irrespective of the year and season. The communities half way the flow path were often also closely resembling those at the start, but in particular after the dry season in 2007 the DGGE profiles revealed a few additional bands. The community profiles close and in the river were in general more different from earlier in the flow path, even though they shared some bands with those samples; bamA genes were found throughout time and space but restriction analysis indicated low diversity in anaerobic aromate degraders. Laboratory culturing experiments showed that microbes were able to remove phenol with an efficiency of up to 70% within 60 days incubation time under sulfate and nitrate reduction conditions. Field data showed that only sulfate was available in the leachate for whole year, thus, degradation of phenol or organic compounds under sulfate reducing condition potentially occurs.

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In Chapter 5, another waste dump was studied: the Keputih landfill close to the sea at Surabaya, Indonesia. This landfill has not been in operation since 2001. Field measurements showed that the leachate electrical conductivity, COD and NH4

+ concentrations were high, containing HCO3-, SO4

2-, Fe2+ and Cl-. Most compounds were again significantly influenced by the season, except COD and SO4

2-. The prime focus was on contamination of the subsurface. Groundwater compositions in almost all monitoring wells surrounding the landfill were indicative of groundwater mixed with landfill leachate. Chemical analysis also suggested the infiltration of seawater. The abundant SO4

2- provides a large potential for anaerobic biodegradation of organic compounds in groundwater. Chapter 6 provides a general discussion and the conclusions on research chapters 2 to 5. Generally, the concentration of organics and ammonium were higher than Indonesian national standards for wastewater Hydrochemistry of leachate from the Jatibarang and the Keputih landfill were strongly influenced by the season; seasonal effects are critical factor in the development of a leachate management policy. Management should be aimed at preventing contamination of groundwater, streams and rivers. The challenge is complicated by the enormous seasonal variation in rainfall, causing widely varying surface and subsurface run-offs. In conclusion, the microbial potential for natural attenuation is certainly present at the landfills, despite a high load of waste materials. Optimal use must be made of this potential, taking into account seasonally varying driving factors. Scientific knowledge of microbe-catalyzed transformations and degradation pathways can help in developing optimal strategies to prevent that leachates cause pollution of groundwater and the environment.

133

Hidro-kimia dan atenuasi alamiah lindi dari tempat pembuangan akhir (TPA) tropis

Ringkasan Penimbunan, merupakan suatu sistem pembuangan sampah kota, yang masih banyak digunakan di daerah tropis, seringkali tanpa perlindungan yang tepat terhadap pencemaran lingkungan sekitarnya. Lindi yang dihasilkan dari tempat pembuangan sampah, dan komposisi kimianya, bervariasi dan sangat dipengaruhi oleh geometri TPA, jenis dan umur sampah, curah hujan musiman, kelembaban dan suhu. Di daerah tropis, kondisi iklim sangat bervariasi dengan musim dan perlu diperhitungkan sehubungan dengan manajemen lindi. Kombinasi proses fisik, kimia dan mikroba dapat memberikan kontribusi pada atenuasi pencemaran di TPA, tetapi pemahaman lebih lanjut tentang proses ini sangatlah penting dalam rangka menyusun manajemen TPA yang lebih baik. Penelitian ini fokus pada beberapa aspek: (a) karakteristik lindi TPA, perubahan musiman dan dampaknya pada pencemaran air permukaan dan air tanah sekitar TPA, (b) transformasi nitrogen yang di mediasi oleh mikroba dan degradasi fenol dalam lindi TPA dan (c) perubahan secara musiman profil komunitas mikroba lindi TPA dan evaluasi kontribusi mereka terhadap atenuasi alamiah. Bab 1 memberikan pengantar umum tentang pengelolaan limbah padat, dengan penekanan khusus pada praktik pengelolaan sampah di Indonesia, dan hubungannya dengan aspek-aspek sosial. Juga, menggambarkan karakteristik umum lindi, dan parameter hidro-kimia yang sering digunakan untuk karakterisasi dan evaluasi. Aspek mikrobiologis dari lindi yang dihasilkan dan proses atenuasi, terutama transformasi nitrogen dan biodegradasi senyawa fenolik, disajikan juga. Lokasi penelitian, pertanyaan-pertanyaan ilmiah dan strategi metodologis disajikan juga dalam bab ini. Dalam Bab 2, Aliran permukaanTPA dari TPA Jatibarang, dekat Semarang, dan dampaknya terhadap sungai terdekat telah dipelajari untuk memahami karakteristik musiman dari lindi TPA dan untuk menentukan terjadinya atenuasi alamiah. Potensi untuk berlangsungnya proses atenuasi alamiah dan kontribusi biodegradasi sepanjang jalur aliran sungai ditentukan dengan cara mengukur secara bulanan berbagai parameter lindi seperti kandungan bahan organik dan komponen-redoks terkait. Hasil penelitian menunjukkan bahwa komposisi lindi sangat dipengaruhi oleh musim. Konduktivitas listrik dan konsentrasi BOD, COD, nitrogen organik, amonia, sulfat dan kalsium selama musim kemarau lebih tinggi (hingga 2,3 kali lipat) dibandingkan pada musim hujan. Waktu tinggal limbah-air dalam sistem penampungan lindi setempat kurang dari 70 hari. Pengukuran lapangan menunjukkan bahwa selama periode ini hampir tidak ada biodegradasi bahan organik dan amonia terjadi (kurang dari 25%). Namun, potensi

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biodegradasi bahan organik dan pengurangan amonia baru jelas terungkapkan selama 700 hari ketika lindi di inkubasi di laboratorium (lebih dari 65%). Tidak ada dampak signifikan dari lindi TPA terhadap kualitas air sungai yang diamati, namun, kehadiran logam berat dan bakteri coliform perlu dipertimbangkan dalam jangka panjang. Pengenceran adalah proses atenuasi alamiah utama yang berlangsung pada lindi. Namun, itu bukan mekanisme atenuasi yang berlangsung secara optimal karena pengenceran tidak dapat mengurangi beban mutlak dari bahan pencemar yang dibuang ke sungai. Amonium adalah salah satu senyawa beracun utama dan pencemar jangka panjang penting dalam lindi TPA. Lindi dari TPA Jatibarang mengandung amonium pada kisaran konsentrasi antara 376 - 929 mg N L-1. Perbedaan besar dalam fluks nitrogen antara musim hujan dan kemarau memberikan kesulitan tersendiri dalam mengendalikan proses pengurangan nitrogen secara biologis dalam lindi TPA. Pemahaman yang benar tentang variasi musiman pada potensi mineralisasi nitrogen, nitrifikasi, denitrifikasi, dan anammox di lindi dapat membantu dalam meningkatkan pengelolaan TPA daerah tropis, dan untuk mencegah penurunan kualitas air permukaan akibat pembuangan yang tidak terkontrol dari lindi yang mengandung konsentrasi amonium tinggi dan beracun. Oleh karena itu, potensi transformasi nitrogen ditentukan dalam Bab 3. Sampel musiman dari kolam pengolahan lindi di jatibarang, diambil lalu digunakan sebagai inokulum diberi makanan berupa media sintetik dan kemudian dievaluasi potensi transformasi nitrogen di laboratorium. Potensi oksidasi amonium secara aerobik (<0,06 mg N L-1 h-1) seratus kali lebih rendah dari amonifikasi dan proses transformasi nitrogen anaerobik, yang dari urutannya sama besar. Pengaruh musim hanya terlihat pada nitrifikasi aerobik dan anammox, dan itu relatif kecil: nilainya mencapai tiga kali lebih tinggi pada musim kemarau. Secara keseluruhan, potensi nitrifikasi dalam lindi TPA Jatibarang adalah yang terendah dari semua proses transformasi nitrogen. Untuk meningkatkan kualitas lindi, dan untuk memungkinkan konsentrasi nitrogen anorganik dalam pembuangan lindi dapat diterima jika dibuang ke sungai Kreo, perlu menerapkan dua langkah sistem aerobik/ anaerobik serta perlu dipertimbangkan untuk dimaanfaatkan lebih optimal. Pengurangan nitrogen secara biologis melalui proses dua-langkah nitrifikasi dan denitrifikasi biasanya disarankan untuk remediasi lindi TPA yang mengandung konsentrasi amonium tinggi. Proses atenuasi alamiah bergantung pada kemampuan kolektif mikroorganisme untuk mendegradasi bahan pencemar di bawah kondisi lingkungan yang berlaku. Biodegradasi adalah mekanisme utama yang paling sering terjadi untuk menghancurkan kontaminan dalam akuifer tercemar. Oleh karena itu, Bab 4 bertujuan untuk mendapatkan wawasan tentang dinamika dan keanekaragaman komunitas mikroba yang terdapat dalam lindi TPA serta potensi mereka dalam mendegradasi senyawa organik beracun, seperti fenol. Pengetahuan ini akan membantu dalam memahami, memprediksi dan meningkatkan proses mikroba yang terlibat dalam proses atenuasi alamiah. Variasi temporal dan musiman pada komunitas mikroba di Jatibarang dipelajari dengan menggunakan teknik denaturating gradient gel electrophoresis (DGGE) profiling, dalam kaitannya dengan hidro-kimia lindi. Kehadiran bakteri anaerobik pendegradasi secara aromatik ditentukan dengan pendekatan kultivasi-independen PCR di mana keberadaan gen (bamA) dari jalur benzoil-CoA degradasi anaerobik ditentukan, dan potensi yang sebenarnya untuk

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degradasi fenol secara aerobik dan anaerobik dengan pendekatan kultur. Antara musim kemarau tahun 2004 dan musim hujan tahun 2006 keragaman yang diamati tinggi pada hampir semua titik sampling, secara umum lebih dari 20 band terlihat dalam profil DGGE. Profil komunitas mikroba berbeda antara titik waktu, meskipun kalau dicermati profil DGGE secara seksama menunjukkan bahwa profil komunitas di lokasi pengambilan sampel tertentu tidak sepenuhnya berubah dari waktu ke waktu, beberapa band tetap ada. Namun, antara musim hujan tahun 2006 dan musim kemarau pada tahun 2007 keragaman mikroba tiba-tiba turun drastis. Sepanjang jalur aliran dari TPA ke sungai diamati terjadi pergantian struktur komunitas. Lokasi sampling di awal jalur aliran permukaan dari TPA ke sungai sering menunjukkan kesamaan yang tinggi dalam profil komunitas, terlepas dari tahun dan musim. Setengah dari komunitas pada jalur aliran sering juga sangat mirip dengan yang berada di awal jalur aliran, tetapi khususnya setelah musim kemarau tahun 2007 profil DGGE mengungkapkan bahwa terdapat beberapa tambahan band. Profil komunitas yang berdekatan dan yang berada di sungai pada umumnya lebih berbeda dari sebelumnya di jalur aliran, meskipun mereka berbagi beberapa band, gen bamA ditemukan di seluruh lokasi dan waktu sampling tetapi analisis restiksi menunjukkan keragaman bakteri anaerobik pendegradasi senyara aromatik rendah. Percobaan kultur secara laboratoris menunjukkan bahwa mikroba mampu mengurangi fenol dengan efisiensi hingga lebih dari 70% dalam waktu 60 ketika diinkubasi dalam kondisi reduksi sulfat dan nitrat. Data lapangan menunjukkan bahwa sulfat dalam lindi tersedia sepanjang tahun, dengan demikian, degradasi senyawa fenol atau organik dalam kondisi sulfat reduksi berpotensi terjadi. Dalam Bab 5, dipelajari pada TPA yang lain: TPA Keputih teletak dekat dengan laut di Surabaya, Indonesia. TPA ini sudah tidak beroperasi sejak tahun 2001. Pengukuran di lapangan menunjukkan bahwa konduktivitas listrik lindi, konsentrasi COD dan NH4

+

berada dalam kategori tinggi, mengandung HCO3-, SO42-, Fe2 + dan Cl-. Sebagian besar

keberadaan senyawa tersebut dipengaruhi oleh musim secara signifikan, tetapi tidak untuk COD dan SO4

2-. Fokus utama dari penelitian ini adalah pada kontaminasi dari sub-permukaan. Air tanah di hampir semua sumur pantau di sekitar TPA terindikasi telah tercampur dengan lindi TPA. Hasil analisa kimiawi juga menyarankan telah terjadi infiltrasi air laut. SO4

2- berada dalam kondisi yang melimpah sehingga berpotensi besar untuk berlangsungnya biodegradasi anaerobik senyawa organik dalam air tanah. Bab 6 menyajikan pembahasan umum dan kesimpulan pada bab penelitian 2 sampai 5. Secara umum, konsentrasi organik dan amonium yang lebih tinggi dari standar nasional Indonesia untuk air limbah. Hidro-kimia lindi TPA Jatibarang dan Keputih sangat dipengaruhi oleh musim, efek musiman adalah faktor penting dalam pengembangan kebijakan manajemen lindi. Manajemen harus ditujukan untuk mencegah kontaminasi air tanah, aliran air dan sungai. Tantangannya menjadi rumit oleh variasi curah hujan musiman yang besar, menyebabkan menyebabkan aliran air permukaanpun sangat beragam Dalam kesimpulannya, potensi mikroba untuk atenuasi alamiah dapat berlangsung di TPA, meskipun beban sampah sangat tinggi. Potensi yang ada harus di manfaatkan secara optimal dengan mempertimbangkan variasi musiman sebagai faktor penentu.

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Pengetahuan ilmiah tentang transformasi mikroba-katalis dan jalur degradasinya dapat membantu dalam mengembangkan strategi yang optimal untuk mencegah lindi sebagai penyebab pencemaran air tanah dan lingkungan.

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Acknowledgements

This dissertation would not have been possible if there had not been support and help from many people. I owe my gratitude to all those people who have made this dissertation possible. I apologize if I do not mention their names one by one here, however, if there is anyone who has contributed, reads this book, but is not mentioned, my gratitude is also addressed to him or her. I did not walk alone in seeking this knowledge, but was guided by two experienced senior scientists. Therefore, to take this opportunity; I wish to thank, first and foremost my promoter, Professor Dr. Nico van Straalen for giving me the opportunity to do my study at VU and for his continuous encouragement and invaluable suggestions during this work. I am not able to express my thanks to him in sufficient words. Many things I have learnt from him, especially his scientific knowledge management and scientific philosophy of evolution. That inspires my career as a teacher and researcher; I hope that one day I will be as good an advisor to my students as Nico has been to me. With immense gratitude I acknowledge the support and help from my co-promoter Dr. Wilfred Röling. His thoughtful guidance, constructive advice, criticisms of this study have greatly supported my Ph.D dissertation toward success. I will never forget that he took the time amid tight activities just to visit the sampling location at Jatibarang landfill in Semarang, Central Java, Indonesia. From him, I have learnt not only knowledge concerning my research, but also a scientific attitude (accuracy, criticism, curiosity and doubt). I don’t have sufficient words to express my gratitude to him. However, I will remember him during the rest of my life. I would like to express my deepest appreciation to Dr. Boris van Breukelen who contributed as co-author to my published articles, and also for making available time to teach me a several concepts on hydrology and the mechanisms of seawater intrusion into groundwater. Without his guidance and persistent help this dissertation would not have been possible. This dissertation becomes more meaningful because of the contribution of a reading committee composed of competent experts. Therefore, I would like to thank Dr. Boris van Breukelen, Dr. Rob van Spanning, Dr. Henk Lubberding, Dr. Boran Kartel and Dr. Anniet Laverman for their constructive comments. I thank Désirée Hoonhout for incredible assistance in the compilation and lay-out of my thesis; I thank Janine Mariën for preparing the cover design. I also learned many skills and techniques from those who are experienced in molecular microbial ecology. I thank Martin Braster for laboratory techniques and providing chemicals. Many important things, tricks and tips I learnt from his

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incredible experiences. I am grateful that during the process of completing this study I made friends in the Department of Molecular Cell Physiology: Bin, Traian, Raquel, Susana, Maria, Christina, Rini and Annachiara. Thanks for sharing experiences; my life was made more colorful with the diversity of culture that I experienced. My special thanks go to Mangihot Goeltom for sharing the Keputih landfill leachate data. I appreciate very much this collaboration, I hope someday I will have the opportunity to take part in your scientific research.

This study was done not due to my personal interest only, but also due to the interests of the institution where I am employed. Therefore, on this occasion, I would like to express my gratitude to the Rector of Universitas Kristen Satya Wacana (UKSW), Prof. Pdt. John Titaley Th.D and his deputies, for giving me an opportunity and financial support to finish my study. Also, I am grateful to the Faculty of Biology, UKSW, especially Dr. Ferry F. Karwur, Dr. Agna Krave and Dr. Rully Adi Nugroho, for making my PhD study possible, and allowing me to work at the laboratory. Also thanks to all staff and laboratory assistants in the Faculty of Biology for collaboration. I saw many unique and interesting things while completing the study, especially from those who emotionally connected to me. Thanks to the experience and knowledge of social solidarity I never felt alone in doing the study even though I had to leave my house and go almost 5000 miles to Amsterdam. Therefore, I thank Karen Vink and Marco Heijkoop for support, especially in last six months when I was in Amsterdam. I will never forget our friendships since I met both of you in Salatiga about 20 years ago. My special thanks go to the Mailuhu-Poerba family for supporting me during my stays in Amsterdam and for accepting me as part of their family. I will never forget all people at PERKI Amsterdam, PERKI Buitenveldert and PERKI Amsterdam Zuid-Oost, for being friends during my work in Amsterdam. From all of you I have learned regarding social life in The Netherlands and through our interactions my Dutch improved. It is nice to remember these moments. I am indebted to all my colleagues in Salatiga especialy to Om John Titaley and Tante Ida, Ferry Karwur and Marry, Dharma and Ina, Roy and Mamy, Theo and Sherly, Om Marthen and Mbak Titi, Rocky and Ling, Yafet and Intan, tante Martha Dandels, mbak Indi, Ronald, Ones Hihika, Sam Papilaya, Ogy, Mikel, Danny and Fanny for your attention, help and support to my sons and especially to my wife, when she was in the hospital. I cannot express my gratitude by words, but I keep it at the bottom of my heart. I thank Gereja Mahesi Injili Halmahera (GMIH) for support and help. I acknowledge all friends in Pusat Studi Kawasan Timur Indonesia (PSKTI), Budhi Prasetyo, Johan Tamboto, mas Eri Budiono, mas Supri, mas Hanung, mbak Dian, and all my students for their support and technical assistance in the field. The success of my PhD studies has to be attributed to the support, help, and prayers from my mother and father, the Mangimbulude family and the Hartoko family.

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My deepest gratitudes go to my wife (Titi Pemata) and my sons (Gereeth and Gregore) for their unlimited love and support throughout my life. I remember their constant involvement and prayers when I encountered difficulties. I feel deeply indepted to them, more than they know. Finally, I like to use this moment to express my gratitude to God for allowing me to learn and “dance” with His invisible creations (the Microbial World). Really! they are amazing creatures, those that stay and live under extreme conditions such as in landfill leachate. Thank God.

Jubhar