LCA Model Landfill

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Waste Management 27 (2007) S58–S74

Landfill modelling in LCA – A contribution based on empirical data

Gudrun Obersteiner *, Erwin Binner, Peter Mostbauer, Stefan Salhofer

Institute of Waste Management, Department Water Atmosphere Environment, BOKU University of Natural Resources and Applied Life Science,

Muthgasse 107, 1190 Vienna, Austria

Accepted 16 February 2007Available online 11 April 2007

Abstract

Landfills at various stages of development, depending on their age and location, can be found throughout Europe. The type of facil-ities goes from uncontrolled dumpsites to highly engineered facilities with leachate and gas management. In addition, some landfills aredesigned to receive untreated waste, while others can receive incineration residues (MSWI) or residues after mechanical biological treat-ment (MBT).

Dimension, type and duration of the emissions from landfills depend on the quality of the disposed waste, the technical design, andthe location of the landfill. Environmental impacts are produced by the leachate (heavy metals, organic loading), emissions into the air(CH4, hydrocarbons, halogenated hydrocarbons) and from the energy or fuel requirements for the operation of the landfill (SO2 andNOx from the production of electricity from fossil fuels).

To include landfilling in an life-cycle assessment (LCA) approach entails several methodological questions (multi-input process, site-specific influence, time dependency). Additionally, no experiences are available with regard to mid-term behaviour (decades) for the rel-atively new types of landfill (MBT landfill, landfill for residues from MSWI). The present paper focuses on two main issues concerningmodelling of landfills in LCA:

Firstly, it is an acknowledged fact that emissions from landfills may prevail for a very long time, often thousands of years or longer.The choice of time frame in the LCA of landfilling may therefore clearly affect the results. Secondly, the reliability of results obtainedthrough a life-cycle assessment depends on the availability and quality of Life Cycle Inventory (LCI) data. Therefore the choice of thegeneral approach, using multi-input inventory tool versus empirical results, may also influence the results.

In this paper the different approaches concerning time horizon and LCI will be introduced and discussed. In the application of empir-ical results, the presence of data gaps may limit the inclusion of several impact categories and therefore affect the results obtained by thestudy. For this reason, every effort has been made to provide high-quality empirical LCI data for landfills in Central Europe.� 2007 Elsevier Ltd. All rights reserved.

1. Introduction

In recent years waste managers, planners and localauthorities have been faced with increasing demands todeliver a sustainable approach to waste management andto integrate strategies capable of producing the best practi-cable option for the environment. Waste managementplanning has become an institutionalised element in public

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* Corresponding author. Tel.: +43 1 3189900 318; fax: +43 1 3189900350.

E-mail address: [email protected] (G. Obersteiner).

planning efforts in all EU Member States. One importantaspect of waste management plans is to ensure the identifi-cation of areas in which specific measures should be takento reduce the environmental impacts of waste management.To demonstrate the performance of management alterna-tives in the decision-making process, authorities, communi-ties, industry and waste management companies shouldconsider environmental aspects in addition to the evalua-tion of technical and economic aspects. It is accepted thatlife-cycle assessment (LCA) concepts and techniques pro-vide solid waste planners and decision makers with anexcellent framework to evaluate MSW managementstrategies.

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The first LCA studies to include the whole life cycle of aproduct in environmental assessment were introduced inthe 1990s. Life-cycle assessment was initially developedfor the purpose of analysing products, although in themeantime it has also been applied to different services.One of these services is represented by the treatment of aparticular amount of waste. However, it should be borneto mind that conventional LCAs were designed to assessproducts rather than services.

In the meantime, waste management evolved to a sepa-rate section within LCA. Since 1998 the InternationalExpert Group on Life Cycle Assessment for IntegratedWaste Management has dealt with this topic (Colemanet al., 2003). Software tools have been created dealing withwaste management issues alone (e.g., ORWARE (Erikssonet al., 2002), LCA-LAND (Nielsen and Hauschild, 1998),IWM2 (McDougall et al., 2001), WISARD (http://www.ecobalance.com/uk_wisard.php)) or as a main issuelike Ecoinvent 2000 (Frischknecht and Rebitzer, 2005).

Initially, LCA experts dealt chiefly with waste manage-ment problems and focused on methodological questions.Nowadays an increasing number of people working inthe waste management sector use life-cycle assessment asmethodology to support decision making. The first groupis clearly involved in methodological problems resultingfrom life-cycle assessment of waste management processes,whilst the second is interested more in specific waste man-agement application.

The use of LCA for resource and waste managementissues implies a slightly different focus than traditionalproduct-oriented LCAs. Most product LCAs do not con-sider end-of-life phases or assume a simplified form of dis-posal. Despite intensive research on this topic during recentyears, the development of a specific modelling of the fate ofthe substances contained in waste disposed of throughincineration, landfill or recycling, is, in view of the currentstate of scientific knowledge, not an easy task. With partic-ular regard to the landfilling of waste or residues of muni-cipal solid waste incineration (MSWI), respectively, or tomechanical biological treatment (MBT), two major chal-lenges are posed by the use of LCA: data gaps and method-ological decisions.

Data pertaining to waste collection, recycling and treat-ment – all processes where direct measurements are possi-ble – are, on the whole, more reliable than data fromlandfills which partially have to be modelled and whereestimations are necessary. A long-range perspective makesexperiments and field studies on landfills difficult to per-form and therefore the uncertainty with landfill models isconsiderable. On top of this is the fact that over the lastdecade landfill design has changed. In the future, accordingto national and EU legislation (landfill directive) furtherimprovements are expected. With more modern wastemanagement strategies including up-to-date treatmenttechnologies such as MBT or MSWI, new types of landfillwill have to be faced. At present due to different reasons,there is a lack of high quality data for these types of land-

fill. In addition to problems of data generation, methodo-logical issues have to be solved, particularly with regardto the time factor where different ideas as to which periodof time emissions from landfill should be considered areunder discussion (Finnveden, 1999). In the mentioned soft-ware models, landfill is included. However, these landfillmodels are based on a number of assumptions and predic-tions about future processes in a landfill. Therefore,in somestudies only the operational stage of waste treatment plantsis considered.

Several authors rightly insist on the importance of theburdens from landfills within the disposal process chain(Hellweg et al., 2003; Doka, 2003; Finnveden, 1999; Sundq-vist, 1999). In practice there is still no panacea on how todeal with the problems emerging when modelling landfillswithin LCA of waste management strategies.

The authors are aware that since the late 1990s in theLCA community, the mentioned problems for LCA inwaste management have been under discussion and manyauthors have published their views on this topic (e.g.,Sundqvist et al., 1997; Finnveden, 1999; Bjarnadottiret al., 2002; Doka, 2003). In this paper therefore the focusis set on providing waste managers with a commented sum-mary of the ongoing discussion from a waste managementpoint of view.

In the past years different multi-input inventory toolshave been developed, most of which are based on theresults of empirical data. Especially in the case of futureMBT landfills, there is a scarcity of sound data. Often land-fill emission data from the literature are used in a ratherarbitrary way. Accordingly, the authors wish secondly topoint out relevant points which should be taken intoaccount when using the literature data and thirdly high-quality empirical LCI data for landfills in Central Europeare provided.

2. Methodological considerations

2.1. LCA in waste management

In talking about life-cycle assessment of integrated solidwaste management, one can distinguish numerous strategicissues which may require different methodologicalsolutions:

1. LCA of a specific product (whole product life cycle orjust waste management system for a given material):the whole product life cycle as well as the waste treat-ment after product use is included within the systemboundaries.

2. LCAs evaluating waste treatment options: assumingthat all upstream impacts are equal, the life cycle ofwaste starts when products are disposed of in the trashbin and ends when the waste material is degraded orreturned to the technological system through recyclingto replace other products. The system boundaries canbe set where the waste is introduced into the system.

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3. LCAs to optimise waste management strategies: Inmany areas this case is similar to alternative 2 but thefocus is placed on the effect of waste management strat-egies. For instance therefore waste prevention may beincluded, implying the need to enlarge the systemboundaries.

Therefore, an LCA focusing on waste is based on differ-ent premises with regard to system boundaries and datacollection.

2.2. Choice of time frame

When applying an LCA approach for waste manage-ment, major data gaps emerge for the long-term emissionsfrom landfills, namely for emissions into water. The presentlevel of knowledge does not allow secure statements to bemade concerning the long-term amount and quality ofleachate either for MBT or for MSWI landfills (Soyez,2001). Existing emission monitoring systems for landfillsare not designed to monitor emitted loads, but mostlyreduced to periodic analysis of the leachate or groundwatercomposition.

Waste in landfills will have a considerably long-lastingimpact on the environment. Several authors including Sab-bas et al. (1998), Sundqvist et al. (1997), Hellweg (2000),and Doka and Hischier (2005) conclude that the pollutantpotential remaining in a landfill 100 years after waste place-ment is significant. In general the environmental potentialof landfills depends largely on the specific material or prod-uct and the type of landfill under investigation. Heavy met-als for example tend to concentrate in landfills and arewashed out to a varying degree over time. Unger (2005)identified in this context three major groups of materials:

� materials with a lower environmental impact in theextraction and processing stage than in the use and dis-posal stage;� materials with a higher environmental impact in the

extraction and processing stage than in the use and dis-posal stage; and� materials with an equal distribution of their environ-

mental impact across the whole life cycle.

Nevertheless long-term emissions may represent animportant burden from landfills. Doka (2004) showed thatthe relevance of the disposal (including incineration and/orlandfilling) may reach 18% of the entire burden from pro-duction and disposal in the case of PVC. On the otherhand, especially for the newer types of landfills (MBT-land-fill or MSWI landfill), in the case of assessment of entirewaste management systems (including transport, collectionrecycling (incl. credits), treatment and landfill), the rele-vance of landfills to the overall results is decreasingremarkably (Wassermann et al., 2005).

During the other stages in the lifetime of a product,emissions will occur more or less instantaneously or at least

within a limited time period. Emissions from other treat-ment methods such as incineration and composting alsocause emissions that occur more or less immediately. Land-fill experts agree that landfills cannot be regarded as stablesystems as to long-term emissions (e.g., Lechner, 2001;Sabbas et al., 1998). The designated barrier systems –inertisation, solidification, sealing sheets etc. – deterioratein time and have a limited functional lifetime. There is arelevant and plausible potential that the remaining pollu-tant load in a landfill will be released completely if longenough time spans are considered (Doka, 2003).

Over the years in general three different phases for land-fills can be distinguished (Sabbas et al., 1999). The short-time phase is characterised by the emplacement of wasteand active and passive maintenance. The time horizonamounts to decades. In the medium-term phase there isno more active maintenance. External factors, environmen-tal impacts on the landfill body are more or less constant.The time horizon comes to centuries. These two phasescan be concluded as ‘‘foreseeable’’. In the long-term phase,where the time horizon reaches 104–105 years, the externalfactors change. Developments are not foreseeable in detail.

The challenge to be dealt with is to select an appropriatetime interval and the time dependent emission function tobe integrated over the selected time interval. In order tomake the potential emissions from landfilling comparableto other emissions during the life cycle, potential emissionshave to be integrated over a certain time period (Finnve-den, 1999). When time frames are discussed, there is aconsensus that the emissions should be integrated over aso-called foreseeable period. However, the time frame forthe foreseeable period may vary from 15 years to 50,000or even 100,000 years in different studies (Sundqvist,2002). The Society of Environmental Toxicology andChemistry (SETAC) recommends that the emissionsshould be integrated over an infinite time period; if this isnot possible, a time interval of 100 years should be applied.In the Ecoinvent 2000 data base (Frischknecht et al., 2004),long-term emissions are defined as emissions occurring 100years after present. They are reported in separate emissioncategories. The landfill inventories include long-term emis-sions with a time horizon of 60,000 years after present(Doka, 2003).

The quantification of the long-term (>100 years)impacts from landfilling is a problem that has not yet beensolved. In the absence of knowledge, some LCAs assumethat these emissions are zero, appearing as a better choicethan other treatment and disposal options. The toxicityimpacts are also frequently neglected because of insufficientscience-based knowledge, making the assessments of haz-ardous waste treatment options difficult. On this accountseveral studies have located the waste disposal outside thesystem boundaries, reporting only the amount of wasteleaving the system boundaries but making no mention ofthe quality of the waste (Koblmuller et al., 2004;Bjarnadottir et al., 2002). An extremely smart solutionfor the time problem was identified by Melloni et al.


(2003) who defined their functional unit as 1 ton of wastesent to the landfill and lying in place for 30 years. The selec-tion of time interval is however, amongst others, an ethicalquestion based on the fact that by limiting the exposuretime, effects on future generations will be avoided. Landfillspostpone the release of emissions from today’s wastes intothe future (Doka, 2005); therefore discounting future bur-dens in LCA studies introduces the risk of burden shiftingfor the future (Hellweg et al., 2003).

When taking into account the general developmentsoccurring over the last 100 years concerning, e.g., technicalinnovations, the prediction of landfill development overexceptionally long time spans is vague at best. Beside fur-ther possible changes of the waste input itself, it dependson many uncertain parameters such as the developmentof geochemical weathering, climate conditions or vegeta-tion (Doka, 2003). The concept of time frames leads tothe necessity of not only describing the outputs but alsothe impact as a function of time. It is however impossibleto ascertain the future characterisation values which mustbe applied to emissions in order to obtain the impactpotential (Van der Ven, 1997).

Thus, when establishing the proper time horizon for theanalysis of environmental burdens from landfills, ethicalissues strongly dictate the demand for adequate data whichstand up to scientific plausibility tests. It would not how-ever always appear to be necessary to give much thoughtto the time frame of an investigation. First, the matter ofsystem boundaries must be settled. As to issues such ascomparison of recycling, incineration and landfilling asalternatives, the choice of the time horizon may representa most important factor in providing results for other ques-tions like treatment of municipal solid waste in MBT,incineration or direct landfilling as alternatives thoselong-term emissions no longer have any effect. The latteris due to the fact that for different landfill types the amountof pollutants released to water and air differs in time. Thesedifferences however only play a significant role in the first100 years.

In sanitary landfills after the methane phase, pollutantscontinue to be released via leachate, although usually on alower level. Emissions to air are negligible. After 100 yearsthe leachate concentrations are assumed to decrease to thelevel of bottom ash landfills (Doka, 2003). There are, ofcourse, distinct differences between the chemistry in a bot-tom ash landfill and a 100-year old sanitary landfill, forexample the content of organic carbon. Nevertheless, inabsence of better data, e.g., the Econinvent database, bot-tom ash landfill concentrations are considered representa-tive once the municipal waste has stabilised to a certainextent and is no longer biologically reactive (Doka,2003). The same assumptions can be made for MBT land-fills: on comparing the leachate concentration of differentlandfill types in the first years and 30 years after landfillconstruction one can find hints referring to this. For exam-ple, during the first years considerable differences arerevealed in concentrations of pH-value and COD (chemical

oxygen demand), as well as for heavy metals. These differ-ences decrease in time. It can therefore be concluded that ifdifferent waste management strategies resulting in differentlandfill types are compared for long-term emissions (over100 years), the amount of waste remaining in the landfillis the only crucial factor pertinent to discrepancies in envi-ronmental pollution.

2.3. Data source: modelling versus empirical results

In general, two different approaches are applied in themodelling of waste disposal processes.

Empirical results from measurements obtained fromvarious landfills of different ages can be used whereverbasic conditions are similar to the specific landfill. Dataare determined in a ‘‘black-box’’ principle with, e.g., muni-cipal solid waste or incineration residues or mechanicalbiological pre-treated waste as input and emissions as out-put incapable of reflecting changes in waste composition.

Multi-input inventory tools on the other hand take intoaccount the initial waste composition. The causal relationbetween the specific waste input and the resulting emissionsmust be calculated by some kind of model. Some of theseemissions are directly dependent on the chemical composi-tion of the product studied. It should however be kept inmind that some emissions are very much process-depen-dent and are therefore difficult to predict.

Within the multi-input inventory tools differentapproaches can be used for modelling:

� To consider the theoretical maximum load, one candefine a maximum emission potential caused by the totalpollutant content in the waste.� To calculate the behaviour of waste in a waste-specific

manner, the degradability of waste fractions is usedand release factors are introduced to portray the re-pre-cipitation of degraded material within the landfill. Lab-oratory tests to determine the availability of substancesembodied in the waste are used.� Another possibility is to carry out model calculation

depending on landfill specific parameters (e.g., transportmodel or geochemical model). Different release factorsare calculated for each chemical element and are cali-brated according to field measurements. Uncertaintiesin chemical and biological interactions and the preferen-tial flow of leachate through the landfill body makesprojections difficult but may also be considered. Onlycalibrated models should be applied.

Fig. 1 provides a review of methodologies currentlyapplied to estimate the environmental loads caused bylandfills. The methodologies vary as to starting material,represented both by data from specific landfills, as well asdata from landfill simulation reactors (LSR). In some casesan attempt was made to alter the starting material, e.g., byevaluating weathering processes as in the beneath describedMPWLP-test (Sabbas and Lechner, 2001).

Fig. 1. Approaches for estimation of loads; T = Time horizon.


In the following the different possibilities will be intro-duced in brief and their pros and cons discussed. It shouldbe mentioned at the outset that in general the multi-inputinventory tools are to be preferred because of the need tomeet the goal of mass balance. The characteristic of wasteas a mixture of various materials adds another element ofcomplexity to the use of LCA. Due to the mixed and var-iable composition of waste, it may be difficult to determinewhich materials in waste cause a given emission (http://waste.eionet.eu.int/lca). For example there should be noemissions of heavy metals from the landfilling of untreatedwood.

If empirical results are used for calculation one shouldbe aware that LCAs require the acquisition of significantamounts of data and that the quality of data determinesthe utility of the final LCA (Bjarnadottir et al., 2002).After studying several case studies and databases, Finnve-den (1998) concludes that data gaps limit the inclusion ofseveral impact categories or cause them to be less coveredand therefore limit the types of conclusions that can bedrawn from these studies. Human and eco-toxicologicalimpact categories are characterised by severe data gapsdue to the large number of possible pollutants in thewaste or produced by waste treatment and to the lackof knowledge of the behaviour of all these pollutants.One means of simplifying the process of data generationis to use legal limit values as maximum possible outputfrom the landfill. This approach can be used at least forcountries where appropriate legislation regulates theamount of pollutants permitted in ground water or airby establishing legal limit values. However, similar legallimit values do not exist for all pollutants and indeed, val-ues are only relevant for the first decades of the landfill.During this time there is a presence of strong reactivity,enabling limits to be exceeded. The technical barrier sys-tem as well as the leachate treatment of the landfill for

instance, is assumed to function properly so that legal val-ues can be met.

Studies performed to investigate the prediction of themid- to long-term behaviour of landfills were carried outparticularly during the last two decades. Many are basedon results gained from laboratory experiments, landfill sim-ulation reactors (LSR) with an experimental volumebetween 0.1 and 1 m3 of waste. Calculating the liquid(leachate)–solid-ratio (LS-ratio) enables the results to berelated to full-scale landfills if well stabilised material isused. This methodology indirectly implies a similar mois-ture distribution in full-scale landfills and LSR (Doberlet al., 2003). Variations in the LS ratio within the landfillbody caused by the heterogeneity of waste, the heteroge-neous character of the landfill produced by compactionprocesses and resulting preferred water paths within thelandfill body are neglected. Doberl et al. (2003) suggestedthat leachate concentration is merely a result of the areassurrounding the preferential flow paths. The degree of het-erogeneity together with the leachate amount determinesthe discharged substance load.

The most detailed remarks on the approaches and back-grounds for calculations of multi-input inventory tools aregiven in Doka, 2003. The Ecoinvent database refers, as faras possible, to the specific chemical composition of thewaste material (waste-specific burdens). In general inmulti-input–multi-output inventory tools, leachate emis-sions from specific elements in landfills (Ee) can be calcu-lated as a function of the mass of the element e in wastefraction (me), the waste specific decomposition or degrada-bility rate (D) and the element specific release factor (re), asthere is discrepancy between the amount of waste that isdecomposed and actual emission from the landfill (e.g.,Doka, 2003; Bjarnadottir et al., 2002).

Ee ¼ m�eD�re ð1Þ

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Additionally, the amount of gaseous emissions from thespecific element must be subtracted. Data concerning thewaste specific decomposition rate (D) can be derived fromthe literature (e.g., Micales and Skog, 1997; Zimmermannet al., 1996; Lechner, 2004). The lack of exact knowledgeof the transfer mechanisms during waste treatment makesit necessary to apply different assumptions to allocate emis-sions to inputs. The crucial point in modelling is the ele-ment specific release factor. Doka, 2003 in the mostrecent Life Cycle Inventory for landfills resorts to empiricaldata by calculating the factor based on the mean annualleachate output from generic landfill data. Of course, Dokais aware of the fact that landfill sites are very heterogeneousregarding the composition of waste and development of thelandfill depending on various conditions (e.g., climate).Although this coarse estimate in general should providetolerable results, not only the developer but also all usersof similar multi-input–multi-output inventory tools shouldbe aware that several elements may be misjudged and thatmoreover, all assumptions are based on specific empiricaldata that may vary as to general conditions (rainfall, tem-perature, age of the landfill when measurements were con-ducted, height of the landfill, waste composition etc.).Currently, only chemical elements such as copper, zinc,and nitrogen are taken into account by disposal modelsas Ecoinvent 2000. Chemical compounds such as dioxinsor other hydrocarbons have not been included to date.The question is whether the methodologies actually appliedfor modelling are generally applicable to the whole periodictable of elements of chemical compounds at all.

An alternative method to model the complex heteroge-neous environment of a landfill is introduced by Zacharofand Butler (2004). As the microbial processes govern thebreakdown of the waste and the hydrological processescontrol the water and leachate movement, they couple asimplified microbial degradation model with a stochastichydrological and contaminant transport model.

In view of the data gap, particularly with concern for long-term emissions from landfills, one has to keep in mind thatthe goal precision of leaching test studies to predict long-term emissions is higher than one magnitude. The Instituteof Waste Management at BOKU-University developed acombined multi-phase weathering and leaching procedure(MPWLP test) for the expanded basic characterisation ofthe leaching behaviour of inorganic waste (Sabbas, 2001).Under the MPWLP-procedure, mineralogical changes thatsignificantly influence heavy metal emissions, especially car-bonation, are accelerated in an artificial laboratory weather-ing step prior to leaching (Mostbauer et al., 2003).

All multi-input inventory tools present the common fac-tor of presuming the disposal of specific single waste frac-tions (not just average waste). Among the many areas ofwaste management, waste disposal represents only a partof the entire waste management system. The exact compo-sition of the waste is not known. Local waste managementsystems and local resource management systems generallydisplay specific characteristics such as composition and ori-

gin of raw materials, emissions from landfills or incinera-tion plants, or energy sources used for electricity supply.Specific data on waste composition are dependent on localwaste management systems. For example separate collec-tion of plastic may include all plastic packaging includingcomposites or may include plastic bottles only. Dependingon the local system the separate collection of a specificwaste fraction therefore does not allow conclusions to bedrawn as to the composition of residual waste or collectedfraction. In order to obtain specific data on waste compo-sition, sorting analyses should be carried out. Dependingon the composition and origin of raw materials, the chem-ical composition of waste materials may also vary greatly.However, in the increasingly globalised economies in whichproducts and waste are imported and exported, the collec-tion of such specific information is difficult and time-con-suming (http://waste.eionet.eu.int/lca). In the highlydeveloped western European or northern American coun-tries, the drawing of conclusions from other countriesmay be acceptable because of similar threshold values aswell as roughly similar external factors such as climate con-ditions. With regard to developing countries or new EUmember states, the characterisation of waste as a mixtureof various materials and therefore the generation of wastespecific data may be difficult.

One should be aware that the respective mobilisation(re-precipitation) rate of individual elements depends onthe amount of newly accumulated leachate, as well as onthe bulk density or the height of the landfill. Calculationsand prognosis can only be made for a specific leachate gen-eration rate and therefore a specific climate and a specificheight of the landfill. Existing tools for waste managementincluding landfills in general do not always allow for con-sideration of specific landfill circumstances such as climate,average rainfall, height and type of cover layer, density,and permeability.

Especially in the case of landfills where the specific pro-cess conditions are often unknown and cannot be con-trolled, it seems to be impossible to develop predictivemodels. This is due to the fact that landfilling is not a singleprocess but rather a series of independent processes (Vander Ven, 1997). Under these circumstances, the model can-not provide an extensive reproduction of reality. The com-plexity of landfills necessitates a multitude of assumptionsand restrictions in modelling. This may be one reasonwhy at present no multi-input inventory tools for MBTlandfills have been developed.

It may be concluded that neither general methodology(empirical results and modelling) can be preferred as theonly ‘‘right’’ solution. One must be aware that both meth-odologies may underestimate or overestimate the actual orfuture concentrations of leachate in a landfill. Thereforethe waste composition and associated maximum possiblepollution should be borne to mind. As already determinedfor the choice of time frame, before opting for one specificmethodology the goal of the study and the practicability ofthe methodology should be taken into account. Different



decisions may be taken for the comparison of differentwaste management strategies or the life-cycle assessmentof one specific product.

3. System characterisation

As discussed in the previous chapters, the providing of agood database for landfills is a very difficult task. There-fore, in the following database several restrictions shouldbe pointed out. The main socio-environmental impactsarising from landfilling can be summarised as followed:

1. the emission of greenhouse, flammable and noxiousgases,

2. the leaching of nutrients, heavy metals and other toxiccompounds with the potential to pollute surface andgroundwater,

3. the generation of noise, dust, odours and other vectors,4. change of the natural scenery,5. the potential to cause deleterious health effects within

local communities,6. an increase in road traffic.

The discussion in this paper only affects the first twopoints of this list, as these are the only points to causemethodological problems dealing with multi-input andmulti-output processes and emissions occurring over a verylong period of time.

When taking into consideration developments over thelast 100 years, predictions for the next millennia seem to beparticularly uncertain. Therefore, although the authors areaware of the fact that within this time period the remainingwaste will generally not be inert, in the following only a timehorizon of 100 years will be taken into consideration.

3.1. Characterisation of analysed landfill types

A further restriction has been made with regard to thetype of landfills addressed in this paper. Only non-hazard-

Table 1Former and actual Landfill types in Europe, Austria, Germany and Switzerla

EU Austria Germany Switzerland

Open dumps Open dumps Open dumps Open dumps

Sanitary landfills Sanitary landfills Sanitarylandfills

Sanitary land

Landfills for hazardouswaste

– Landfill classIII

Landfills for non-hazardous waste

Massenabfall-deponie

Landfill classII

–

Reststoff-deponie Landfill classI

Residual matlandfills

Landfills for inert waste Bodenaushub-deponie

Landfill class0

Inert materialandfills

Baurestmassen-deponie

ous waste landfills are taken into account. Table 1 pro-vides an overview of landfill classification in theEuropean Union, Austria, Germany and Switzerlandand their allocation to each other. In addition to currentlandfill types, former landfills that are no longer author-ised according to the new EU Directive 1999/31/EC orcountry specific legislation but which still exist, are cited.For the eco-inventory the grey coloured landfill types wereincluded.

Already in Austria and many EU member states and inother EU member states in the near future landfills mayonly accept waste that has either been pre-treated byincineration in order to attain a TOC of less than 5%or has undergone mechanical biological treatment. Obvi-ously former landfill types such as sanitary landfills andopen dumps present as old deposits in Austria as wellas in Europe, or which may even still be installed in someparts of Europe should be included in the followingconsiderations.

To include landfilling in an LCA approach entails sev-eral methodological questions as shown in Section 2.Additionally, for the relatively new types of landfill(MBT landfill, landfill for bottom ash from MSWI) noexperiences are available for the mid-term behaviour (dec-ades). Thus, for our analysis, a review of the literaturewas performed to evaluate both data from lab-scale testsas well as larger scale tests. Additional experience anddata from our own investigations of MBT landfills andresults from MPWLP tests (Binner, 2003; Mostbaueret al., 2003; Wurz, 1999) were included. Only referenceswith proper meta information were used for the LifeCycle Inventory.

3.1.1. Landfilling of untreated waste

When municipal solid waste is landfilled directly, anaer-obic biological degradation produces landfill gas and leach-ate. Over 90% of the converted organic carbon is releasedas CO2 and CH4. The remainder is released in the leachate(Binner, 2003).

nd

Explanation

Landfills for untreated municipal waste, without specifictechnical safety measures

fills Landfills for untreated municipal waste, with physical barriers toprotect the public from waste

Landfills for hazardous waste

Landfills for MBT waste

erial Landfill for MSWI residues

l Landfill for excavated earth

Landfill for construction and demolition waste


3.1.2. Impact of the mechanical biological pre-treatment on

landfill behaviour

In Austria it was possible to gain valuable knowledgeabout the long-term behaviour of wastes that had beenmechanically and biologically treated (Binner, 2003).Impacts on the landfill volume as well as on the water bud-get, the leachate composition and the landfill gas emissionare known. The treated waste had a placement density of1.3 ton/m3, saving 30–50% of landfill volume. If sieve res-idues are incinerated, calculations showed that it is possi-ble to save up to 77% of landfill space (Raninger, 1995).The water budget of a landfill is strongly influenced bylocal conditions. However, investigations showed thatbecause of a remarkably low permeability only a verylow amount of leachate can be expected. Fehrer (2002)observed extremely low permeability in two different sam-ples of pre-treated wastes (kf < 10�10 m/s resp. kf < 10�11

m/s). Leachate analyses proved that the acidification phaseis eliminated by mechanical biological pre-treated waste(Binner, 2003). Leachate concentrations are in a lowerrange of concentrations than normally found in reactortype landfills. Binner (2003) concludes that MBT landfillsshow 90–95% less organic burden than other landfills ofthe same age filled with untreated waste. Heavy metal con-centrations were also reduced. The duration of pre-treat-ment has a very significant impact on the generation oflandfill gas. Compared to conventional landfilling ofuntreated waste, a reduction of gas production of up to95% can be reached.

3.1.3. Impact of municipal waste incineration on landfill

behaviour

After municipal waste incineration, the remaining resi-dues shall not have more than 3%wt. of organic carbon(TOC). In contrast to landfills of untreated municipalwaste among MSWI landfills, low landfill gas formationcan be expected because of the low organic content. Eventhough one cannot exclude gas generation entirely, becauseof the small database, no statements about amount andtime dependent behaviour of gaseous emissions fromMSWI landfills can be made (Kabbe, 2000).

Decisive for the release and immobilisation of fixedheavy metals in incineration residues is the developmentof pH value and redox potential of the landfill body. Typi-cally the pH-value from the leachate from MSWI slag in thefirst decades after deposit is about 10–11. Because of reac-tions between the alkaline components of the slag with car-bon dioxide from the air (or from landfill gas) and due toother weathering processes, the pH-value will decrease tovalues between 7.0 and 8.5. At the same time the LS ratioincreases. In total a decrease (‘‘wash out’’) or stabilisationof the annual load (‘‘solubility control’’) of inorganic pollu-tants occurs. If the slag cannot react with carbon dioxide ina sufficient way, e.g., because of a very fast buildup of thelandfill body, a strong mobilisation of aluminium, zincand lead may occur. Altogether in MSWI landfills, a dis-charge of increased pollutant concentrations can be

expected at least occasionally (Hirschmann, 1999). Chemi-cal treatment of leachate is necessary in most cases becauseof the initial alkalinity and heavy metal content.

3.2. Characterisation of environmental impacts of landfills

3.2.1. Landfill leachate

As for the leachate, the following parameters have to beborne in mind: The accumulating amount of leachate for alltypes of landfill depends on the same factors that can besummed up in meteorology, latitude and altitude, vegeta-tion, material properties and morphological factors. Nor-mally the amount of leachate is given as a percentage ofrainfall. It should be considered that the amount of rainfallinfluences the resulting amount of leachate. If the amountof leachate is approx. 50% in regions with 1000 mm ofannual rainfall, this is not the same for regions with500 mm of annual rainfall. In regions with lower rainfall,the amount evaporating on the surface is comparativelyhigher than in regions with higher rainfall.

As additional factors to rainfall and climate (tempera-ture dependency of evaporation), the landfill geometry con-stituted by height and density as well as surface and landfillengineering (placement technology, cover, leachate man-agement system) have to be considered (Binner, 2003). Asan example, reports published for the amount of leachatefor reactor landfills range from 25% to 60% of rainfall(Krumpelbeck, 2000; Plinke et al., 2000; Wallmann,1999). According to Rettenberger and Fricke (1998), atoptimised landfill operation of MBT landfills the amountof leachate may fall below the area of 9–13% of rainfall.

The quality of leachate depends on the toxic potential ofthe landfilled material, the amount of leachate and the ageand related age-dependant environmental factors of thelandfill. The formation of preferential flow paths was mon-itored for the disposal of solidified wastes for different land-fill types. The formation of cracks may occur, e.g., becauseof settlements and later on lead to the formation of waterchannels. The therewith associated changes in the water bal-ance of landfills are difficult to predict. A consequence maybe that water in the remaining, bypassed parts of the systemmust flow considerably slower, or not at all (Rosqvist andDestouni, 2000). One can assume as well that a dilution ofleachate occurs after intensive rain due to this preferentialflow. This fact has to be considered at the interpretationof leachate data.

Due to this considerable variation in leachate, it isimportant that the system studied reflects the actual tech-nology applied in the geographical and temporal scope ofthe study. The description of medium and long-termbehaviour of landfills by means of simple recipe like modelsseems to not be possible. Therefore assumptions have to bemade.

3.2.2. Landfill gas emissions

The predicted amount of gas emissions from landfillscan be estimated by the content of organic matter in waste.

Table 2Data on gas potential in the literature

Source Gas potential Remarks

Stegmann and Dernbach (1982) 150–200 m3/ton DM (approx. 90–120 m3/ton FM) ExperimentalTabasaran and Rettenberger (1987) 120–300 m3/ton FM Laboratory experimentsPlinke et al. (2000) 150–190 m3/ton FMKrumpelbeck (2000) 180–280 m3/ton DM (approx. 100–170 m3/ton FM) Data from different landfills

DM, dry matter; FM, fresh matter.


The direct landfilling of municipal solid waste is expectedto have a gas potential from 100 to 180 m3/ton DM (drymatter) (Table 2). The gas potential related to the disposedwaste (fresh matter, FM) is of importance in calculations.Table 3 gives an overview of data on gas potential foundin the literature. According to investigations carried outby our group on solid municipal waste in the province ofSalzburg, a gas potential of 120 m3/ton FM was used forthe modelling in this paper (Binner et al., 1999).

By providing a suitable mechanical processing of wasteand optimised control of rotting process by aerobic pre-treatment, a decrease of gas potential to about 5–10% basicvalue can be reached (Heiß-Ziegler et al., 2004; Soyez, 2001).Measures on an MBT-landfill in Austria showed a gaspotential of 11.5–21.5 m3/ton (Rolland, 2001). In our study

Table 3Specifications concerning to used landfill references

Source Location Type of material landfilled Da

Krumpelbeck(2000)

The old WestGerman states

High amount of household waste, aswell as construction waste,contaminated soil and sewage sludge

36.gas

Wurz (1999) Attnang-Redlham,Austria,

80% Mechanical biological pretreatedwaste, 10% bulky, demolition andcommercial waste; operation period1980–1983; landfill closed 1983

Loana(Ph(Ph

Turk et al.(1997)

Wilhelmshaven,Germany

MBT-material n.s. n.s

Johnsonet al.(1999)

Lostorf nearBuchs,Switzerland

MSWI bottom ash monofill, filled in1992; not covered yet

AvdurfroNo

Hjelmar(1995)

Vestskoven,Denmark

MSWI ash monofill, containing 10,000tons of bottom ash with 15% fly ash

An197ave(ph2)

Hjelmar(1995)

Woodburn,Oregon US

Combined ash monofill La

Kruse (1994)andKabbe(2000)

Laboratoryvalues

Untreated residual waste DS

Zweifel et al.(1999)

Switzerland,Riet

Mainly household waste, landfilled1925–1935

Loana

Mostbauer(2005)

Rautenweg,Vienna, Austria

Weathered incineration residues;MSWI bottom ash and MSWI bottomash containing some other ashes

Meanafro

n.s., not specified.

according to the RA4 (respiration activity) limit volume ofthe Austrian landfill directive (RA4 < 7 mg O2/g DM) andto our own investigations of pre-treated material, a gaspotential of 15.6 m3/ton FM was calculated (Table 4).

4. Empirical LCI data for different landfill types in Central

Europe

In the following inventory, data for the four differentlandfill types under investigation are presented for the threedifferent time periods. Data are from the literature and ourown investigations. In view of the fact that in leachate com-position, water represents the transport medium for pollu-tants it is important to focus on the sources, factors andprocesses which govern the discharge. Climatic conditions

tabase Comment

780 leachate analysis, 3.970analysis

Evaluation and assessment of 76 landfills;less than half of them with landfill bottomliner, most landfills with leachate and gascollection system; average annual values

ngstanding series of leachatelysis; data from 1980 to 1983ase 1) and 1984 to 1992ase 2)

. Values for AOX, Na and K for Period IIof MBT landfills

erage values from 194 samplesing experimentation periodm November 28, 1994 tovember 15, 1996

Values for average conditions excludingrain and dry periods

nual analysis of leachate3–1992, 22 observations;rage values for 1973/1974ase 1) and 1991/1992 (phase

Values for 0–2 year old landfill and 18–19year old landfill;

ndfill age about 2–7 years Only maximum values for Na, K, Cafrom Woodburn

R values n.s.

ngstanding series of leachatelysis; data from 1994 to 1997an value from leachatelysis between 1999 and 2003

m Rautenweg test cells;

Compare Jaros and Huber (1997), Sabbaset al. (2001) and Mostbauer (2005) fordetails about the laboratory based workin the field of weathering of MSWIbottom ash

Table 4Summary of assumptions for landfill models

Open dump Sanitary landfill MBT-landfill MSWI-landfill

Annual landfill rate (t) 242,958/104,483a 104,483 29,315 26,762Average landfill height 10 mBulk density 1 ton/m3 1 ton/m3 1.4 ton/m3 1.5 ton/m3

Bottom layer No State of the art State of the art State of the artCover layer No Methane oxidation layer Methane oxidation layer Recultivation layer

Operation period (I) 5 yearsActive aftercare period (II) 30 yearsModelling period (III) 100 years

Landfill gas potential 120 m3/ton FM 120 m3/ton FM 15.6 m3/ton FM –Period I 22% 22% 30%Period II 75% 75% 60%Period III 3% 3% 10%Active landfill gas collectionPeriod I No No NoPeriod II 45%Period III NoMethane oxidation rate (Periods II and III) No 90% 90% –Average annual rainfall 1300 mm 1300 mm 1300 mm 1300 mmAmount of leachate [% of rainfall]Period I 50% 50% 50% 65%Period II 50% 30% 5%/30%a 30%Period III 50% 30% 5%/30%a 30%Leachate treatmentPeriod I No B-MF-RO RO PF-AC-OxPeriod II B-MF-RO RO PF-AC-OxPeriod III No (impermeability of bottom layer uncertain)

B-MF-RO: Biological treatment, MicroFiltration, Reverse Osmosis.PF-AC-Ox: Precipitation/Flocculation, Activated Carbon filter, Oxidation.

a Two variations considered.


(precipitation characteristics, temperature and wind), vege-tation and the type of surface soil determine the waterbalance.

The discharge pattern also depends on the permeabilityof the material, as well as the possible presence of preferen-tial path flows. Johnson et al. (1998) proved that in the pres-ence of preferential flow paths, the discharge from a MSWIbottom ash landfill is characterised by long periods of lowand nearly constant flows interspersed with increases inleachate quantity after rain.

With a view to landfill gas composition, the cover layerand type of surface, respectively, are particularly importantas determining factors. Engineered methane oxidation cov-ers are very effective in reducing greenhouse gas emissionsfrom landfills. Water retention layers and well-plannedrecultivation at the surface may reduce leachate amountssignificantly.

Therefore, when publishing landfill emission data orworking with landfill data from the literature, it is impor-tant to know as much as possible about the metadata, thebackground conditions responsible for the specific landfillemissions. Beside the general climate conditions and theage of the landfill, it is fundamental to ascertain as far aspossible the generation of data pertaining to leachate (aver-age data over a certain time period, quantity of measure-ments, single analysis during rainy or dry periods, etc.).

Table 3 provides a review of the references used in prepara-tion of the following inventory tables.

In the present paper it is assumed that, with the excep-tion of ‘‘Open Dumps’’, a fully functional leachate treat-ment has been implemented during the first 30 yearsfollowing landfill construction, and that 100% of the leach-ate is collected and treated in this active aftercare period. Inthe opinion of the authors, the impermeability of the bot-tom liner system can no longer be guaranteed after this per-iod, after which direct leachate emissions into groundwaterare to be expected. To consider the leachate emissions intogroundwater, the eventual presence of geological barriershas not been taken into consideration and therefore hasto be taken into account separately.

To take into account different quantities and qualities ofleachate in view of the age of the landfill, emissions wereconsidered separately for the periods 0–5 years (operationperiod), 6–30 years (active aftercare period) and 31–100years (medium time period, where active aftercare hasstopped). All assumptions for modelling are summarisedin Table 4. After a detailed literature survey, values weretaken only from studies where detailed meta-informationconcerning the landfills and generated data was given(Table 3).

Tables 5–10 provide detailed information on the lifecycle inventory of different types of landfills. They are

Table 5Leachate concentration for different landfill types during operational period I (year 1–5); in cases when values reported in the literature exceeded legal limitvalues and legal limit values were used, these values are marked in grey


based on a black box concept, meaning that depending onthe amount of waste landfilled, the specific emissions canbe calculated from the concentration of the specific elementin the leachate or gas and the amount of leachate or gasdeveloped.

In view of the fact that for sanitary landfill, MBT landfilland MSWI landfill leachate, treatment is obligatorythroughout the first 30 years, the specific emission valueshave been replaced by threshold values according to theAustrian sewage emission ordinance (‘‘Abwasseremissions-verordnung’’, BGBl 263/2003; BGBl 179/1991; BGBl.Nr.186/1996) in cases when values reported in the literatureexceeded this legal limit value. These values are markedin grey (Tables 5–7).

Additionally, expenditures incurred in leachate treat-ment (e.g., electricity) should be calculated, being identicalfor both landfill operation or for the use of energy fromcollected landfill gas. These figures should be calculatedaccording to the specific conditions under investigation.

Tables 5–7 show the concentration of pollutants in theleachate of the four different landfill types investigated. Inorder to avoid any attempt at obtaining a non-existent

degree of accuracy, values are rounded. Empty fields indi-cate that no useable values were found. Zero indicatesthat measured values are below limits of detection. Tocalculate the leachate emission potential for the first 100years, the leachate flow rate according to Eq. (2) shouldbe applied

Eleach; 100 ¼XððRf � Leach%I � cLeach; IÞ

þ ðRf � Leach%II � cLeach; IIÞþ ðRf � Leach%III � cLeach IIIÞÞ ð2Þ

Eleach, 100 is the leachate emission for element e in the first100 years of landfill, Rf is average annual rainfall (ml),Leach%I is amount of leachate in percent of the rainfallin periods I, II or III from Table 4 (%), cLeach I to III is leach-ate concentration of the element e for Periods I, II or IIIfrom Tables 5,6 or 7.

Tables 8–10 show the concentration of pollutants inthe leachate of the four different landfill types investi-gated. The process of calculation is slightly different, start-ing from the overall landfill gas potential. To calculate thegas emission potential for periods II and III (year 6–100),

Table 6Leachate concentration for different landfill types during active aftercare period II (year 6–30); in cases when values reported in the literature exceededlegal limit values and legal limit values were used, these values are marked in grey


active gas collection and methane oxidation should alsobe taken into consideration, as illustrated in Tables 9and 10.

5. Summary and conclusions

Not only the production and the use of materials canproduce a burden to the environment but also the treat-ment and disposal of the same products as waste providesa far from negligible contribution to environmental pollu-tion. It is therefore common sense that the disposal phaseshould by default be included in LCA studies. Both froma waste management and a life-cycle assessment point ofview it is clear that disposal, and therefore the emissionsoccurring from landfills, should be included in the evalua-tion of waste management options as well as in the assess-ment of a products life cycle.

As discussed in the present paper the above consider-ation leads to several methodological questions. The mainquestions debated are related to the issues of time frameand data generation. It is common sense that the use oflong time frames (infinite, if possible) and the use of

calculation methodologies taking into account the specificwaste composition should be applied. Both aspects leadto time-consuming and therefore costly procedures.

In contrast to other processes involved in the productionor management of waste where the emissions from the pro-cess can be measured directly and immediately, this is nottrue for landfills. Landfill emissions do not occur at onepoint (like a smokestack) but there are diffuse emissionsmanifested over a very long time span. A direct measure-ment of pollutant loads occurring over many decades isnot possible. Therefore assumptions have to be made.

When including landfill emissions in an assessment pro-cedure one is confronted with numerous uncertainties onthe one hand and very often with a complicated calculationprocedure on the other. This paper attempts to demon-strate how the latter depends particularly on the goal andcontent of the specific LCA, and to question whether it isreally necessary or even useful to work with long time‘horizons exceeding the calculable time period of 100 yearsand/or to calculate waste specific emissions by means ofmulti-input–multi-output tools. It is demonstrated thatalthough for many issues such as comparison of recycling,

Table 7Leachate concentration for different landfill types during foreseeable time period III (year 31–100)

Landfill type Open dump Sanitary landfill MBP landfill MSWI landfill

Mean Mean Mean Mean

pH-value 7.70 7.70 7.70 7.60TOC mg/l 500 500 500BOD5 mg/l 300 300 300COD mg/l 1200 1200 1200NH4–N mg/l 120 120 120NO2–N mg/l 0.84 0.84 0.84NO3–N mg/l 10.00 10.00 10SO4 mg/l 80 80 80 2.900SO3 mg/l 1.1 1.1 1Cd mg/l 0.0028 0.0028 0.0028 0.0027Fe mg/l 12.50 12.50 13 0.09Zn mg/l 0.54 0.54 0.54 0.13AOX mg/l 1.13 1.13 1Na mg/l 600 600 600 500K mg/l 600 600 600 200Ca mg/l 200 200 200 600Mg mg/l 100 100 100 400Mn mg/l 0.91 0.91 1 0.07Pb mg/l 0.03 0.03 0.03 0.0027Cu mg/l 0.04 0.04 0.04 0.03Ni mg/l 0.12 0.12 0.12 0.03Cr mg/l 0.18 0.18 0.18 0.01Hg mg/l 0.0004As mg/l 0.04 0.04 0.04 0.0047Al mg/l 0.05Sb mg/l 0.03B mg/l 10.00 10.00 10 3.27Ba mg/l 0.03Co mg/l 0.0017Mo mg/l 0.26Si mg/l 4V mg/l 0.0043

Table 8Landfill gas concentration for different landfill types during operational period I (year 1–5)

Landfill type Open dump Sanitary landfill MBP landfill

Landfill gas potential m3/ton 120 120 15.6Proportion year 1–5 % 22% 22% 30%Landfill gas potential year 1–5 m3/ton 26.4 26.4 4.68CH4 Vol% 60 60 60CO2 Vol% 40 40 40CH4 m3/ton 15.8 15.8 2.8CO2 m3/ton 10.6 10.6 1.9

Min Max Min Max Min Max

CH4 g/t 11,339 11,339 11,339 11,339 2010 2010CO2 g/t 20,738 20,738 20,738 20,738 3676 3676Benzene mg/ton 0.26 2508 0.3 2508 0.9 3.7Toluene mg/ton 0.53 16,236 0.5 16,236 0.1 13.5o-Xylol mg/ton 5.28 185 5.3 185 0.1 12.6p/m Xylol mg/ton 0 9926 0 9926 0.3 32.5Dichloromethane mg/ton 0 6600 0 6600 0.017 0.327Trichloromethane mg/ton 0 52.8 0 52.8 0.002 0.006Tetrachloromethane mg/ton 0 15.8 0 15.8 0.002 0.0021,1,1-Trichloroethane mg/ton 0.013 726.0 0.013 726.0 0.003 0.042Trichloroethene mg/ton 0 4804.8 0 4804.8 0.014 0.050Tetrachloroethene mg/ton 0.003 3748.8 0.003 3748.8 0.005 0.032


Table 9Landfill gas concentration for different landfill types during active aftercare period II (year 6–30)


Landfill gas potential m3/ton 120 120 15.6Proportion year 6–30 % 75% 75% 60%Landfill gas potential year 6–30 m3/ton 90 90 9.36Active gas collection % 45%Gas used for energy production m3/ton 40.5Landfill gas emissions year 6–30 m3/ton 90 49.5 9.36CH4 Vol% 60 60 60CO2 Vol% 40 40 40CH4 m3/ton 54 29.7 5.62CO2 m3/ton % 36 90 19.8 90 3.74Methane oxidation %CH4-emissions m3/ton 54 2.97 0.562CO2-emissions m3/ton 36 46.53 8.798


CH4 g/ton 38,657 38,657 2126 2126 402 402CO2 g/ton 70,698 70,698 91,378 91,378 17,279 17,279Benzene mg/ton 0.9 8550 0.495 4703 1.760 7.413Toluene mg/ton 1.8 55,350 0.99 30,443 0.140 26.947o-Xylol mg/ton 18 630 9.9 347 0.197 25.132p/m Xylol mg/ton 33,840 18,612 0.599 64.977Dichloromethane mg/ton 22,500 12,375 0.034 0.654Trichloromethane mg/ton 180 99 0.004 0.012Tetrachloromethane mg/ton 54 30 0.003 0.0031,1,1-Trichloroethane mg/ton 0.045 2475 0.025 1361 0.007 0.083Trichloroethene mg/ton 16380 9009 0.028 0.099Tetrachloroethene mg/ton 0.009 12,780 0.005 7029 0.010 0.065

Table 10Landfill gas concentration for different landfill types during foreseeable time period III (year 31–100)


Landfill gas potential m3/ton 120 120 15.6Proportion year 30–100 % 3% 3% 10%Landfill gas potential year 6–30 m3/ton 3.6 3.6 1.56CH4 Vol% 60 60 60CO2 Vol% 40 40 40CH4 m3/ton 2.16 2.16 0.936CO2 m3/ton % 0 1.44 90 1.44 90 0.624Methane Oxidation % 0 90 90CH4-emissions m3/ton 2.16 0.216 0.094CO2-emissions m3/ton 1.44 3.384 1.466


CH4 g/ton 1546 1546 155 155 67 67CO2 g/ton 2828 2828 6646 6646 2880 2880Benzene mg/ton 0.036 342 0.036 342 0.293 1.236Toluene mg/ton 0.072 2214 0.072 2214 0.023 4.491o-Xylol mg/ton 0.72 25.2 0.72 25.2 0.033 4.189p/m Xylol mg/ton 0 1354 0 1354 0.100 10.830Dichloromethane mg/ton 0 900 0 900 0.006 0.109Trichloromethane mg/ton 0 7.2 0 7.2 0.001 0.002Tetrachloromethane mg/ton 0 2.16 0 2.16 0.001 0.0011,1,1-Trichloroethane mg/ton 0.002 99 0.002 99 0.001 0.014Trichloroethene mg/ton 0 655 0 655 0.005 0.017Tetrachloroethene mg/ton 0.0004 511 0.0004 511 0.002 0.011


incineration and landfilling as alternatives the choice of thetime horizon may represent a most important factor, withregard to other issues such as treatment of municipal solidwaste in MBT, incineration or direct landfilling as alterna-tives such long-term emissions no longer have any effect.

This is due mainly to the reason that for different landfilltypes the amount of pollutants released into water andair differs according to time. These differences howeveronly play a significant role in the first 100 years. The sameapplies for the choice of methodology, where the decision


should be made on the basis of the goal of the study as wellas on the resulting system boundaries. When carrying out aproduct LCA we usually wish to establish some kind ofcausal relations between a specific product or materialand the emissions actually caused by the product studied(Sundqvist et al., 1997). If a waste management system(starting from waste generation not from production) isthe object of investigation, even raw data of waste compo-sition may be lacking and indeed at times be superfluous incases where the same type of waste is treated and landfilledin different ways. The option of considering the mostextreme case (all substances contained in deposited wastewill be released from the landfill sooner or later) in whichthe chemical composition of waste is known, constitutesan easy to use alternative.

Further to performing a detailed examination of meth-odological issues the second goal of this paper was toprovide a useful database of landfill emissions for the cen-tral European region. The inventory compiled from emis-sion data from four different landfill types for threedifferent periods of time provides a detailed data basefor direct use wherever general conditions apply to theconditions described in the paper, as well as for furthercalculations for universally valid modelling e.g. forMBT landfills.

A considerable challenge to the use of LCA in waste andresource management is represented by the fact that theimpact of these systems depends largely on regional condi-tions, including consumer habits, mode of transport, gener-ation of by-products and energy, or the energy supplysystems in place (fossil fuels, biomass, hydropower,nuclear, wind). It is therefore important that when possiblelocal data be applied, rather than importing external dataor using the default data.

As a conclusion one can identify three different areaswhere the focus of further investigations should be laidon. As outlined in the paper one big problem for the assess-ment of the impact of landfills within LCA is still the lackof data. Both specific emission data as well as correspond-ing meta data have to be investigated in long-term studies.Special interest should be given to chemical compoundssuch as dioxins which often are not included into existingdatabases. This is mainly due to the fact that the contentof those compounds in leachate is often near the limit ofdetection and therefore a calibration of models is difficultor even impossible. But not to include those substancesin the LCA may affect the result because of their e.g. highhuman toxicity potential.

In general in the last decades a lot of models have beendeveloped. For the future it seems necessary that suchmodels have to be subject to a sort of quality assurancewhich includes: (a) verification of the assumptions andthe program code, (b) calibration with field data or datafrom large pilot plants and (c) external evaluation.

As a third area of interest one has to detect the relevanceof consideration of landfills in LCA-analysis of disposalsystems for different scenarios to give a guideline for fur-

ther investigations under which conditions landfills haveto be included and in which degree of detail.

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