14.2.1 Clinker, cement and concrete containing wastes

The use of potentially pollutant recycled materials can be observed at different stages in the processing of cement and cement-based materials.

Portland clinker is obtained at high temperature (1450 °C) from limestone, clay and a combustible mixture defined to obtain a fixed composition of Ca, Si, Al and Fe oxides. The high energy demand of the process is partially satisfied by the replacement of fossil combustibles by waste materials, a substitution that is also called co-processing or co-incineration. Recycled materials with convenient characteristics for the combustion process and for the clinker quality are used: high calorific power, composition, size distribution, humidity, etc. The fossil combustible substitution rate is between 14% and 90% of the total energy demand (ADEME, 2009), impacting the different factories economically and environmentally. Traditionally the substitution combustibles (called alternative fuels and raw materials, AFR) are composed of high caloric residues (waste oils, tyres, plastics, etc.), dried sewage sludge, and solid residual fuels (SRF) containing plastics, wood, paper, cartons, etc. The control of AFR composition is highly important and specific operations are necessary to prepare it (GTZ-Holcim, 2006). They are typically fed to the kiln system via the raw material supply in the high-temperature zone. The clinker reactions allow the incorporation of ashes and binding of metals to the clinker. Statistically, the effect of AFR on the heavy metal composition of clinker is marginal, except for Zn content in the case of bulk use of tyres (GTZ-Holcim, 2006). The use of substitution combustibles in the cement rotary kilns does satisfy precise compositional constraints for high-quality clinker, e.g. sulphur, chlorine (SRF), heavy metals (waste oils, SRF, etc.) or phosphorus content (meat and bone meal, sewage sludge, etc.).

Commercial cements are mixes of fine crushed clinker with different mineral admixtures designed to improve their specific properties, e.g. to improve the pozzolanic properties or the setting and/or to reduce price. To clinker are added gypsum (up to 5 wt%) for setting regulation and different other secondary materials with pozzolanic action: granulated blast furnace slag from iron metallurgy, coal combustion fly ashes, burned shale, calcareous limestone, silica fumes. The composition of the common European commercial cements is detailed in the EN 197–1 standard, e.g., CEM I (Portland cement) is composed of 95% clinker and up to 5% secondary constituents, while CEM VA contains clinker (40–64 wt%), furnace slag (18–30 wt%), pozzolanes and siliceous fly ashes and up to 5 wt% secondary constituents. Compositions of special cements are also standardised, e.g. cements to be used in marine environments or in sulphate-rich waters. The characteristics of these admixtures are consequences of precisely defined limits for the cement properties, e.g. the insoluble residue, the S and Cl content, the pozzolanity, etc. In the composition of these admixtures, specific parameters, like their heavy metal content, are subject to investigation because of their potential environmental impact.

The granulometry of the aggregates used in cement-based materials divided them traditionally into mortars (granulometry less than 4 mm) and concrete. The concrete formula is correlated with the utilisation scenario: mechanical and environmental requirements, planned lifetime, conditions and constraints of implementation. This defines the composition to be ensured: content and type of cement, additives, size classes and composition of aggregates, water content, and fabrication protocol. The economic and environmental benefits of replacing more or less totally the natural aggregates by aggregates with similar properties from the reuse of materials like inert waste, demolition concrete, residues from different thermic processes like municipal waste incineration, slags and sands from metallurgical processes, are obvious. The feasibility of the substitution requires not only good physical parameters (such as particle size, density, colour, etc.), chemical and mineralogical composition of the aggregates but also easy implementation, risk and traceability control.

Concrete admixtures are obtained from secondary materials like lignin or issue from specific chemical synthesis. Strictly speaking, they are not all recycled materials, but their long-term life cycle in concrete can lead to their degradation to potentially hazardous products. Plasticisers, super-plasticisers, air-entraining admixtures, waterproofing compounds, retarders, setting and hardening accelerators, grouting aids and stabilisers, defoaming agents, etc. are used because of their beneficial effects during concrete processing; workability and setting behaviour can be better controlled. The properties of the hardened concrete (strength, impermeability, etc.) can also be improved by adding additives. Therefore, important technological, economic and energetic (and thus environmental) benefits are achieved by using these additives. It is estimated that more than 90% of commercial concretes are produced with admixtures, the majority of them being plasti-cisers/superplasticisers, basically composed of modified lignosulfonates, melamine–formaldehyde sulfonate polymers, naphthalene–formaldehyde polymers, naphthalenesulfonate, etc. Generally 80-90% of the initial additive mass is sorbed by the hydrated cement phases. The unsorbed part of the admixtures remains in the pore solution and can be subject to coupled transport and reaction (degradation) processes and subsequent hazardous effects on the environment and human health.

14.2.2 Structural materials and aggregates

Reuse/recycling of wastes to substitute and/or complement natural aggregates is mostly encountered in road construction and other earthworks (embankments, backfilling of mines, quarries, excavations, etc.).

A common road structure is composed of several layers of different natural aggregates with or without binders. The embankment comprises at least two layers realised with coarse aggregates at the bottom, and fine aggregates used alone and with a hydraulic binder (lime) at the top. The sub-base and the base layers confer mechanical resistance and require aggregates consolidated by hydraulic binders or bitumen. Finally, the pavement layers are generally composed of aggregates and asphalt. The wastes suitable for use in road construction are those providing a good mechanical resistance, compacting property and binding capacity. The wastes most often employed are the municipal solid waste incineration bottom ashes (MSWI-BA), fly ashes from coal combustion, slags from metallurgy, waste concrete, mining wastes, and foundry sands.

For waste use in backfillings the same general characteristics are requested as for embankments. In addition, lightness and ease of pumping are useful properties for such operations, and could be ensured by coal fly ashes.

14.2.3 Main waste streams used in construction materials

The European statistics on waste production and management are of poor quality. Some data on production and reuse/recycling in the construction sectors of the main waste streams in Europe in recent years are presented in Table 14.1.

MSWI-BA residues represent about 80% of the incineration residues of municipal solid wastes and are considered to be non-hazardous. Their chemical composition depends on the municipal waste composition and incineration process. The main components are silica, lime, alumina, limestone, traces of heavy metals and unburned organic matter. The material could be reactive in contact with water, releasing soluble salts and evolving through more stable solid phases. A pre-treatment of the fresh waste allows improving the chemical stability by separation of different fractions and by ageing (hydration–carbonation) for immobilisation of certain pollutants.

The material is suitable as a secondary aggregate, mainly used for road embankments, foundations and backfilling, but also for minor applications like sound-insulating walls, breeze blocks, or non-structural safety-related purposes such as concrete and bituminous concrete aggregate (Abbott et al., 2003; data for eight EU countries, estimations for the period 1990–2000).

Coal combustion for heat and electricity production generates several coal combustion products (CCP) depending on the combustion process and coal origin:

According to ECOBA (2011), production of CCP in 2008 was approximately 56 million tonnes in EU-15 and 100 million tonnes in EU-27, with the following composition: fly ashes (66.6%), FGD Gypsum (20%), bottom ash (8.6%) and boiler slag (2.4%). Applications for CCPs in construction materials are numerous: as additives in concrete, as a cement replacement material, as an aggregate or binder in road construction (good compacting, pumping and binding properties), and as a filler.

Slags are rock-like or glassy materials composed chiefly of lime and silica. Minor components are magnesia and alumina. The exact ratio of these constituents depends on the process at the origin of the slag.

Post-furnace treatment is carried out to enhance and optimise the properties and qualities of slags to ensure that the requirements of the end user are completely matched. Data published by EUROSLAG (2006) for European countries concerning steel and blast furnace slag production by the EU steel industry in 2004 are presented in Table 14.1. These wastes are mainly reused as aggregates and binders for road construction, public works, concrete, mortars and grouts, in cement fabrication, and in other materials like bricks, pavements, breeze blocks, etc.

Construction and demolition wastes represent about 31% of total waste generated in Europe (ETC/SCP, 2009; statistical data from 2008 for EU countries and Norway). Many components in this waste category are easily recyclable and have the potential to replace up to 10% of raw materials. The composition is very variable and changes in time and from one country to another. Data on recycling and reuse in the construction sector are not reliable for many countries. However, a mean EU value for total recycling/ reuse of 46% is reported (EC/DG ENV, 2011; data for the EU-27 countries).

Concrete, bricks and tiles represent together about 78% of construction and demolition wastes. Other components are ceramics, gypsum, glass, metals, asphalt, plastics, treated and untreated wood, dredging soil, track ballast, sand, gravel, stones, etc. In addition, if not separated at source, such waste can contain small amounts of hazardous wastes (insulating foams, paints, asbestos, etc.), the mixture of which can raise particular risks to the environment and can hamper recycling. The mineral fraction represents the great majority of construction and demolition wastes, up to 85%. This mineral part can replace natural aggregates in road construction, earthworks and structural concrete. Asphalt is 100% recyclable in hot or cold mixing, replacing virgin aggregates and binder. Waste treated wood can be reused for timber product fabrication if it complies with the chemical contamination limits (EC/DG ENV, 2011).

Construction and demolition waste has been identified as a priority waste stream by the European Union. In particular, there is a reuse market for aggregate wastes in roads, drainage and other construction projects. In addition, technology for the separation and recovery of construction and demolition waste is well established, readily accessible and in general inexpensive.

Mining, quarrying and ore-processing wastes represent about 27% of the total waste produced in Europe per year (EC-EDG, 2004). Most mining and quarrying wastes are waste rocks (resulting from the excavation of the mine sites) and tailings which are generated at all levels of the recovery process to upgrade the minerals, and are considered as ultimate waste impoverished in useful elements. These wastes are of great chemical variety depending on the natural resource exploited, the chemical species extracted from them and the extraction process used (BRGM, 2001). The main categories related to the extraction process and chemical composition are waste from coal extraction, waste minerals from industrial mineral extraction usable as such after concentration or purification (talc, kaolin, bentonites, etc.), and wastes from metal or other element extraction by physico-chemical separation processes. Coarse mining waste and especially barren rock is sometimes considered as material for roads, building foundations or cement factories, depending on its geotechnical and geochemical characteristics.

Various other wastes can be used in construction materials, such as glass, foundry sands, rubber from tyres, plastics, sewage sludge, etc.

14.3.1 Waste used as alternative fuels and raw (AFR) materials in cement kilns

Depending on the properties of the waste feed, pre-processing could be necessary, such as sorting, neutralisation, drying, etc. The major part of AFR is solid with various shapes, dimensions and densities, so special operations are necessary to grind it optimally to convenient homogeneous dimensions and facilitate metal extraction (magnetic, hydraulic, etc., processes), for meeting the required conditions for kiln feeding. Some other wastes can be liquids (oils and solvents, from chemical, transport or agricultural industries) or slurries. The composition of the AFR and its variability in time can be important parameters influencing the cement quality and its potentially hazardous impact. Some classes of waste are prohibited, like radioactive waste, explosives, electronic waste, batteries and unsorted municipal waste because of their variable properties or high content of heavy metals.

The mineral part of AFR must bring a material value to the cement kiln. Limit values for the content of some elements of AFR are proposed by national regulation or different industrial permits (ADEME, 2009; GTZ-Holcim, 2006), taken individually or by group of elements with close properties (such as total halogens, alkalis, etc.), as a function of the origin of the waste or otherwise. Here are some examples showing quite large differences between the composition of admitted waste/AFR in the cement kiln: Cu (less than 100–700 mg/kg), Pb (less than 150–800 mg/kg), Zn (less than 400–500 mg/kg), Cr (less than 50–500 mg/kg), As (less than 13–20 mg/kg), Co (less than 12–100 mg/kg), Ni (less than 30–200 mg/kg), Cd (less than 2–30 mg/kg), Sb (less than 5–800 mg/kg), Be (less than 2–5 mg/kg), Hg (less than 0.5-10 mg/kg), V (less than 10–100 mg/kg), Sn (less than 10–100 mg/ kg), S (less than 10–400 mg/kg), Cl (less than 0.2%–2%), F + Br + I (less than 0.2%–1%), etc.

The organic part of AFR represents a calorific value for the cement kiln. However, particular toxic organic components are monitored, like PCBs (less than 50 mg/kg to 1%), cyanides (less than 100 mg/kg), benzene, dioxins, etc. The admission of AFR in the high-temperature zone of the kiln should allow the complete degradation of the organic components (exothermic oxidation) and stable incorporation of metals in the clinker phases.

14.3.2 Additives used during the hydration process of cement for mortar and concrete fabrication

As shown before, the different organic additives added during cement processing into mortar or concrete can remain partially sorbed on the hydration phases and therefore in direct contact with the pore solution contained in the porous system. In fact, the pore solution is very concentrated in alkaline species (pH > 13) and can interact with the additive molecules, breaking them down into more soluble and mobile molecules (Spanka and Thielen, 1995). If the initial additives do not have hazardous properties, the molecules obtained by their degradation can be potentially hazardous. Due to their higher mobility they can be leached out from the material in more or less short term. The global mechanism can be diffu-sional, controlled by species release in the leachant. Literature data point out, for example, that lignosulfonates are broken down into by-products like vanillin and eluted, while melaminesulfonates are rapidly broken down into insoluble by-products, less eluted in relatively short-term experiments (Duchesne and Bérubé, 1994). Parts of the generated species are known to be hazardous, potentially carcinogenic pollutants: phenols, formaldehyde, naphthalene, etc. However, the environmental impact of leached organics from concrete (produced by the degradation of the additives) has traditionally been considered insignificant because of the low proportion of additives used in commercial concretes (about 1% of the cement weight) and the small amount of the added quantity that is affected by degradation and release.

14.3.3 Recycled aggregates

The potentially hazardous character of recycled aggregates resides in their chemical composition. Generally, these wastes are composed of a solid matrix that is relatively inert (silicates and oxides stable in contact with water), crystalline (often very porous) or vitreous, which also confers the mechanical properties of the aggregates. Various trace elements are also present depending on the origin of these recycled materials: ores, coals, residues from thermal processes, and municipal wastes (in the case of MSWI-BA). Some of these trace elements can have toxic properties depending on their chemical speciation, solubility in contact with water, and means of exposure. Several examples are given below.

The slags issued from different metallurgical processes represent annually huge quantities potentially available for reuse: e.g. Waelz slag from the recycling of electric arc furnace dusts (weight composition: 7–23% CaO, 4–7% Al₂O₃, 0.1–0.5% Cr₂O₃, 4–40% FeO, 0.01–0.1% As, 0.3–0.5% Cu, 0.4–4.2% Pb, 0.2–4% Zn, 0.8–2% S, etc.) (Barna et al., 2000a), Imperial Smelting Furnace slag (average weight composition: 18.1% CaO, 10.4% Al₂O₃, 33.7% FeO, 0.1% As, 0.7% Pb, 7% Zn, 1.6% S, 21.9% SiO₂, etc.) and Lead Blast Furnace slag (average weight composition: 20% CaO, 1.9% Al₂O₃, 33.4% FeO, 0.15% As, 3.5% Pb, 11.2% Zn, 0.7% S, 23% SiO₂, etc.) from the primary non-ferrous metallurgy (Barna et al., 2004), etc.

MSWI-BA composition is variable depending on the municipal solid waste composition and type of incineration process, the major elements being Si, Ca, Al, Fe, Na, Mg, K and C. Elements like Cl, Ti, Zn, Cu, Ba, S and Mn could be found in concentrations of about 1–10 g/kg. Trace elements (less than 1 g/kg) include Pb, N, Sn, Cr, Zr, F, B, Ni, Sb, V, Co, Cd, Ga, Li, La, Mo, Ba, As, Be, Au, Sc, Hg, Se and cyanide (Jeong et al., 2005). The organic part composed of unburned matter varies up to 30% of the MSWI-BA dry mass.

The chemical composition of fly ashes (CCP) depends on the coal origin and combustion process used. Besides the major constituents (Si, Al, Fe, Ca) a large variety of chemical elements are present in trace quantities, some of them being targeted for their potential hazard. These are Cr, Cu, Ni, Pb, V and Zn present in several hundreds of mg per kg of dry fly ash, As and Se as dozens of mg/kg, and Hg, Cd and Sn up to 1 mg/kg (Rakotoarisoa, 2003). Concerning the chemical speciation, the trace elements are mostly captured in stable alumino-silicates and oxides, reducing their solubility and availability in contact with water. Condensation of volatilised elements during combustion can also occur, especially at the surface of ash fine particles.

Significant European research programmes have been dedicated in recent decades to the characterisation of waste materials in parallel with the development of an environmental assessment methodology (Barna and Blanc, 2011). A huge bibliography is available concerning the leaching properties of MSWI-BA, fly ashes and slags (obtained by laboratory leaching tests, Section 14.4).

Generally, the amount of pollutants released can be linked to their total content. However, their mineralogical characterisation can explain apparently contradictory behaviours, e.g. the encapsulation of small metallic droplets in the glassy phase of slag (5–47 wt%) explains a reduced release of Pb and Zn between the different Waelz slags (Barna et al., 2000a). Leaching conditions such as leachate composition or exposure conditions (carbonation, cycles like wetting/drying, freezing/thawing, etc.) can contribute to particular behaviours. For example, organic compounds having complexation capacity can explain the high level of Cu release in materials containing MSWI-BA (Bröns-Laot et al., 2004), and acidic conditions increase the release from ‘reactive’ materials like phosphates (apatites) or cement-based materials containing pollutants.

14.3.4 Potential toxicity of wastes used in construction materials

In the European Waste Catalogue (EC, 2001), Annex III is dedicated to the hazardous wastes characterised by at least one of the 14 EWC-defined hazardous properties. Criterion H14 defines ecotoxic wastes as ‘substances and preparations which present or may present immediate or delayed risks for one or more sectors of the environment’. The ecotoxic properties of a hazardous waste can be assessed using chemical composition and/or biological methods, realised on solid waste and/or on the leachate (standardised as EN 14735, CEN, 2003). The biological method is based on ecotoxicity tests (bioassays) realised on a set of selected species and has the advantage of integrating all pollutants with their synergistic, antagonistic and additive effects as well as their bioavailability.

The classification of wastes as hazardous is determined by the presence of at least one toxic effect or by at least one concentration higher than a maximal accepted value (standard). The concentration limit values are not consensual in the EU; specific national regulations coexist (e.g. the composition of water destined to potabilisation or landfill acceptability).

The ecotoxic potential of several waste types used in construction has been determined following the prescriptions of EN 14735 and applying different ecotoxicity tests; several results are given here.

MSWI-BA is one of the most investigated wastes. Different studies performed over the last 10 years and in different laboratories confirmed that the incinerator ashes can be hazardous or not, depending on their origin and combustion process. In all cases the leachates have basic pH and high concentrations of Cl^–, SO₄^–2 and dissolved organic carbon.

A set of seven MSWI-BA of different origins were investigated for eco-toxicity potential (Lapa et al., 2002) using bioassays on leachates (with a bacterium, a freshwater alga, a crustacean, and a vegetable). This study showed that the identification and classification of the ecotoxic potential (and finally the decision concerning the waste’s valorisation) depend on the control parameters chosen (chemical composition vs standard limit values, chemical composition vs bioassay, and different bioassays) and pointed out the necessity of a common regulation in this field. In this study, all wastes were found to be ecotoxic according to the chosen standard.

A test battery applied (Römbke et al., 2009) on leachate (algae, Daphnis, and luminescent bacteria) and on solid waste (plants, earthworms and bacteria) showed that globally the aged MSWI-BA is slightly less toxic than the fresh samples. No correlation was found between toxicity and heavy metal concentrations in the leachate and solid waste samples; the toxicity seems to be due to the high salt concentration.

Concerning the ecotoxicity assessment, the method described here lacks clear choices for ecotoxicity tests and standard limit values; as of today, no consensus is established for a mandatory method (Wilke et al., 2008).

14.4.1 Health risks in the construction and demolition stages

A risk for human health could arise directly in the fabrication and demolition stages, when the workers are directly exposed, mainly through inhalation. Fugitive dust emissions occur from stockpiles, deposits and unfinished works. Operations like transport, handling, different processing steps (e.g. sorting, crushing, grinding, shredding, sieving, pulverising, chipping, abrasion, spreading, compacting) generate fine particles which are dispersed in the atmosphere. Whatever the chemical composition, dust causes harmful stress for workers and over long exposure can induce industrial diseases. The harmful effect is amplified if the particles contain toxic compounds.

Another source of health problems is represented by materials based on organic compounds which can be released in the atmosphere by volatilisation at ambient or process temperatures. This could be the case with plasticisers, solvents, glues, etc. existing in waste materials (mainly construction/ demolition wastes). Semi-volatile organic compounds (e.g. aromatic hydrocarbons, dioxins, furans, etc.) can be released during heating in specific processes such as hot processing of asphalt and gas-phase reactions after combustion of organic wastes in cement rotary kilns.

Dust and condensed compounds can be deposited in the neighbourhood of the work site and contaminate soil, vegetation and water resources, and then be transmitted by the food chain to animals and humans. The emission of dust can in certain cases be avoided by water spraying, or by covering or enclosing the machinery; dust inhalation can be avoided by using gas masks.

Another aspect to be considered here is the potential leaching of the particulate materials by rainwater during the fabrication and demolition steps. The leaching phenomenon generates soluble compounds and their dispersion in soils and natural waters, and also the runoff of fine and colloidal particles. The more the material is divided in finer particles, the more the leaching process is amplified and the release of toxic compounds increased. So, if a block-compacted material can be relatively inert, the same material dispersed in small particles can become a real threat.

Very poor information on the harmful effect on workers during the construction/demolition steps, or via the food chain, is available. General aspects such as those discussed above are mentioned by EC/DG ENV (2011) for construction demolition wastes, and by Abbott et al., (2003) for MSWI-BA. These latter authors present a risk assessment study considering a mean scenario of road construction using MSWI-BA as aggregate in asphalt pavement. This study is based on a series of worst-condition hypotheses for the emission of dust (containing pollutants like As, B, Ba, Cd, Cr, Cu, Hg, Mo, Pb, Sb and Zn), for organic volatile compounds (dioxin-like molecules) and also for leaching by rainwater on the road construction site. The dynamic of pollutant concentration in air, water and soil was calculated, and pollutant transmission by the food chain was estimated. Intake by inhalation was considered for the workers and the ingestion route for the local foodstuff consumers. Finally, the exposure parameters for target humans (the hypothetically most exposed individuals) were estimated (means of exposure, concentration, duration and dose), then the predicted doses by pollutant were compared with toxicological effect parameters. For all scenarios considered, the predicted doses were small compared with relevant reference doses for metals and dioxins, and even in the worst-scenario hypothesis the method did not foresee any impact on human health. The authors stress the fact that the study was carried out on a given waste sample and the results could not be extrapolated for all MSWI-BA streams due to the variability of their chemical composition.

14.4.2 The service life – emission scenarios

During the service life of construction materials studied in this chapter, the main hazard potential is represented by contact with water, a common component of the litho-and biosphere. Water is the major pollution vector in the environment; in contact with a material, the soluble/mobile constitution substances are dissolved and carried out through soils and water to the biosphere. The leaching phenomena are thus responsible for mobilising potentially hazardous substances, impacting the quality of soil, water and finally the biosphere and human health via contact and intake routes (water and food). Given the long-term spread and the multitude of situations, this stage will be presented in more detail. The leaching and emission scenarios are defined with respect to the way the construction material comes into contact with water.

A first criterion of classification may be the vector characteristics associated with the conditions of exposure to water (Schiopu et al., 2007): (1) outdoor exposure to water – meteoric, surface and underground water; (2) indoor exposure to vapours and cleaning solutions; (3) exposure to particular water: sea, sulphate reach, etc.; (4) exposure to water intended for human consumption; (5) exposure to wastewater; and possible combinations between these classes. The last two classes concern particular devices for water consumption and treatment, and, with the second class, are marginal to the present topics. In the following, attention is paid to the first class because of the variety of situations and the quantity and type of construction materials involved.

Retaining the first class of outdoor exposure to natural waters, a second criterion could be the type of contact with water: how the water circulates at the material surface and in the material. Four typical scenarios generate different impacts for the same exposed material, as described below (Schiopu, 2007):

A fifth category could be introduced corresponding to works completely immersed in water like dykes, piles and basins made generally of concrete-based materials, granulates, wood, etc.

14.4.3 Parameters of influence

The main mechanisms involved in pollutant emission from construction materials (source terms) are chemical and transport processes. In contact with water, the material undergoes chemical and structural changes. At the liquid/solid interface dissolution/precipitation, adsorption/desorption, and different complex surface biochemical reactions occur, depending on the chemical composition of the material and also on the presence of exogenous chemical species like gases (e.g. atmospheric CO₂ and O₂), and dissolved species existing in the natural water (rain, underground or surface, sea, etc.).

If the material is non-porous (e.g. glass, metallic slabs), only the contact surface is reactive and participates in pollutant emission. If the material is porous – as in the majority of cases – the reactions occur not only at the ‘apparent’ surface but also in the pores, anywhere a liquid phase is present.

The dispersion of the pollutants takes place by different transport mechanisms:

The chemical reactions and transport of dissolved species out of the material are complicated dynamic processes producing structural changes e.g. modification of the porous structure, erosion, modification of the chemical composition of the solid matrix, etc., finally influencing the further release behaviour of the material. Generally, increase of the water/solid contact surface leads to intensification of the chemical interactions and pollutant release. A granular material will release greater quantities than a block of the same mass.

The life span of the material and the contact with water (duration, flow, etc.) are parameters influencing the quantity of pollutant released. The dynamics of pollutant release generally vary over time because the materials themselves evolve, the chemical composition and structure (and implicitly the pollutant speciation) changing. Water contact is generally not continuous (except in situations of immersion) but intermittent, depending on climate. A periodic sequencing of similar events can be observed, e.g. rain periods in a season. The influence of the water contact duration and sequencing on the pollutant release is difficult to foresee and still remains a research and standardisation objective.

The temperature influences the chemical processes (equilibrium constants and kinetics) and also the water’s physical state. Humidification and drying cycles of porous materials play an important role in the evolution of mechanical and chemical structures. Material humidification favours dissolution and diffusion of substances, while drying determines the migration of water from deep pores towards the material surface and precipitation of dissolved species. Phenomena like swelling or cracking are encountered for certain materials. Finally, ageing, which is the modification of the chemical, physical and mechanical structure of the material under the effect of natural factors (time, UV, temperature variations, humidity, etc.) and functioning stresses, influences the pollutant behaviour through the mechanisms described above.

To conclude, all the parameters of influence can be classified as (1) material intrinsic parameters, i.e. chemical composition, porosity, size of the product, state of the surface in contact with water (profile, geometry, roughness) and homogeneity, and (2) scenario parameters, i.e. water composition, contact mode (intermittency, velocity), life span, mechanical conditions, climate conditions, air, water, temperature, humidity, UV, etc.

14.5.1 Methodologies and tools

To evaluate the toxic effects of pollutants for humans and the environment, specific methods and tools must be used. All the methods are based on the general methodologies of human Health Risk Assessment (HRA, described in a series of documents available from national organisations for health surveys (EPA, 2011)) and Ecological Risk Assessment (ERA guide (EPA, 1998)). The main steps of HRA are (1) hazard identification: type of problem the pollutant can cause; (2) exposure assessment: how the target person is exposed to the pollutant in time, at which concentration or dose; (3) dose–response assessment: how the pollutant quantitatively affects the target; and (4) risk characterisation: calculation of the risk of occurrence of harmful effect for the exposed persons. The ERA methodology is based on (1) exposure assessment, (2) assessment of the pollutant effect on living organisms, and (3) risk characterisation.

If HRA steps 1 and 3 and ERA step 2 characterise a given substance with respect to a target organism, the ‘exposure assessment’ step in turn determines the way the pollutant is emitted by the source (in waste containing construction materials in service life scenarios) and transported to the living targets. This implies a good knowledge of the pollution source in terms of emitted pollutant flux and its dynamics over the period of concern and represents the key step for a reliable HRA or ERA.

Often it may be sufficient to determine the exposure model, i.e. the time-dependent pollutant concentration in a given environmental compartment (soil, sediments, groundwater, surface waters, etc.). Then the pollutant concentrations can be compared with acceptable limit values for a given harmful effect. This is the principle of many simplified methodologies proposed and applied in different countries for assessing the eco-compatibility of waste reuse as secondary materials in earthworks and road construction (Sweden, Hartlen et al., 1999; Netherlands, Eikelboom et al., 2001; Denmark, Hjelmar et al., 2001; Norway, Petkovic et al., 2004). The limit values are taken for each hazardous pollutant from toxicology/ecotoxicology databases or are already specified in specific regulations (like the water directive – water destined for human consumption). However, questions remain concerning, for instance, synergy effects between pollutants in complex dynamic release.

The European standardisation organisation has elaborated a methodological approach for the characterisation of the pollution source term in scenario conditions, EN 12 920. The key point is the link between the leaching behaviour of the waste material and the exposure conditions. The methodology proposes an iterative experimental and modelling multistep approach intended to provide predictive up-scaled pollutant releases.

In the following, we focus on the material leaching behaviour as a key point for hazard assessment. The above-mentioned methodologies call for experimental tools in order to determine the intrinsic material properties like chemical and mineralogical composition, porosity and pore size distribution, etc., as well as the scenario parameters and the time evolution of these characteristics. Then the information obtained can be used in models with the aim of calculating missing parameters (not available from experiments), extrapolating, forecasting and simulating different situations.

14.5.2 Leaching and toxicity tests

The information on material/water contact conditions and their influence on potential release of pollutants may be obtained from appropriate leaching experiences at laboratory, pilot or field scales. Many conventional leaching tests have been designed to compare leachate composition to regulatory thresholds: the compliance tests. For example, EN 12457.1 to 4 concern four different but quite close leaching protocols on granular material each during 24 hours. Beyond them, tests dedicated to the ‘characterisation of the leaching behaviour of waste’ have been elaborated for understanding the release mechanisms and provide necessary data for their modelling. The main leaching tests elaborated at the European level are presented below; these tests are expected to be transposed for construction material assessment.

Because of the complexity of waste-containing materials, the use of solubility thermodynamic models for the different species is considered not reliable enough and it is necessary to complete data by experimental means. Two tests have been elaborated to satisfy this objective using similar principles, i.e. to reach a steady state in solid/liquid processes at controlled pH, for different pH values: CENTS 14429, ‘Influence of pH on leaching with initial acid/base addition’ and CEN TS 14997, ‘Influence of pH on leaching with continuous pH-control’. In order to facilitate the achievement of a steady state, specific procedures are foreseen: fine crushing of the material, efficient stirring, negligible atmospheric CO₂ uptake, long contact time, etc. The data (elemental eluate concentrations) collected by these pH-dependent solubility studies are useful in the analysis of the leaching results and are implemented in the modelling process of leaching, including the steps of calculating the local dissolution/precipitation processes at the pore level.

CEN TS 14405, ‘Up-flow percolation test (under specified conditions)’, is specifically designed for the characterisation of pollutant release by percolation of water in a column filled with granular material by means of leachate time-dependent composition analysis. The controlled constant up-flow has to be performed until a cumulative liquid/solid ratio of 10 L/ kg dry matter is reached. However, characterisation of the residence time distribution of the leachant in the system is missing.

The leaching of ‘monolithic’ blocks (at the laboratory scale) allows study of the leaching behaviour in the context of the solid material, at a fixed leachate (water) volume/surface of the sample. The leachates are collected in time and analysed. Two variants exist (CEN TS 15863 and CEN TS 15864) differentiated by the protocol of the leachate renewal (periodic leachant renewal and continuous flow of leachant). These tests allow calculation of the fluxes of released pollutants and supply important input parameters for leaching process modelling.

Although standard experimental tools exist for the leaching assessment of waste and waste-containing materials issued from waste stabilisation/ solidification processes, there are no standard experimental tools for construction materials evaluation so far. Experimental difficulties can be foreseen, like very low supposed pollutant concentrations, the long testing time, the coupling of biophysico-chemical processes, etc. Research studies have been performed in particular cases but it is too early to come to universal conclusions and methods.

Several regulation texts are being drafted concerning the experimental tools dedicated to assessing the release of dangerous substances from construction materials. The Technical Committee TC 351 of CEN (European Committee of Normalization) launched the concept of a horizontal testing procedure (CEN/TR 16098:2010) which might be a common methodology for testing any substances and construction products covered by the Construction Products Directive. At the European level, the CEN/TC 351/ WG1 working group is mandated to study the ‘release from construction products into soil, ground water and surface water’ and to propose the appropriate experimental (horizontal) leaching tests. The emerging leaching tests (and corresponding EU standards) are a monolith leaching test and an up-flow percolation test, very close to those already used for waste characterisation.

The toxic or ecotoxic parameters are usually determined by bioassays – experimental tools for measuring the pollutant effects on living organisms. The bioassays can be applied at different steps in a risk assessment approach. In very upstream steps, the bioassays could be applied to materials in order to estimate an intrinsic toxic/ecotoxic potential. However, the results are relevant only for the laboratory conditions used in the tests. The real toxic/ ecotoxic effect must be determined at the scenario level: in the utilisation conditions influencing the chemical and leaching behaviour of the exposed material and with the target living species. The experimental conditions of the bioassays must therefore be as effective as possible in simulating the real scenario conditions.

As already shown in the presentation of the leaching tests, the construction materials are not covered by the waste regulations. In Section 14.3.4 on waste characterisation, the requirements of the toxic character evaluation are satisfied by an experimental procedure based on ecotoxicity tests, realised on solid waste and on a leachate (obtained by a given protocol). As of today, there is neither regulation nor guidance for how the bioassays must be performed on construction materials. Generally, the eco-toxicity assays can be performed on different media (solids, liquids) using appropriate living species for these media and sensitive to the studied pollutant. The experimental conditions (dilution, temperature, light, culture medium, etc.) are specific for the species used and for the monitored effect (mortality, growth, reproduction, luminescence, etc.). In the case of a chemically complex matrix (solid or liquid), complex effects like synergy, antagonism and bioavailability exist and render the test results difficult to interpret. In the case of construction materials like those discussed in this chapter, typical mixture effects such as many toxic elements (heavy metals, metalloids, extreme pH values, high salinity of leachates) contribute to the bio-assay response. It is therefore difficult to correlate the chemical composition and structure of the material with the ecotoxicity test result. Specific bio-assays are developed for evaluating chronic, acute and genetic toxicity, for different trophic levels: producers (plants, algae), consumers (fish, larvae) and decomposers (bacteria, earthworms). There are numerous standard ecotoxicity tests at both international and national levels elaborated initially for chemical substances, and then developed for different matrices such as industrial wastewaters and polluted soils.

A huge amount of documentation on toxicity and ecotoxicity tests is available on the websites of regulating organisations such as the OECD (Organisation for Economic Co-operation and Development, OECD, 2011), ISO (International Organization for Standardization, ISO, 2011), US EPA (United States Environmental Protection Agency, EPA, 1994), and others. The direct application of these tests on wastes and materials containing wastes, on solid or leachate, raises the issue of the significance of experimental conditions (especially the dilution and the leaching duration) and consequently of the relevance of the ecotoxic result and interpretation. Ideally, the bioassays should be realised on a great variety of species and effects relevant to the scenario conditions.

14.5.3 Modelling

The leaching of hazardous substances from construction materials in a specific real scenario (the exposure assessment) could be evaluated by modelling the physico-chemical (and biological) processes, based on experimental data as mentioned above and on the scenario’s physico-chemical parameters.

Leaching behaviour interpretation and modelling require taking into account different mass transport and chemical phenomena with different complexity levels. The modelling methods employed in practice are various, from empirical to completely mechanistic, with different prediction capabilities. Effective case-adapted models are proposed in the literature, some of them using ‘home’ developed models and software, most being reduced to the main processes and having limited extrapolation capabilities. As the chemistry is very complex because of the chemical nature of the studied materials, high-performance numerical tools coupled with huge thermo-dynamic databases are needed. Such modelling tools for aqueous chemistry are well known as geochemical software, some being used more and more in the field of waste and materials leaching, e.g. PHREEQC, ORCHESTRA, MINTEQA2, CHESS, EQ3/6, WATEQ4F and Geochemist’s workbench. Appropriate software suites coupling chemistry and transport are fewer and contain different transport models (PHREEQC, HYTEC, PHAST) for particular applications.

Figure 14.1 schematises the conceptual structure of a coupled model for leaching behaviour assessment in a real scenario. The modelling procedure should be based on the three main steps: (1) specific hypotheses for chemical and transport processes -mathematical development and model resolution, (2) identification of the unknown model parameters by using laboratory experimental data and (3) model validation by using pilot measured data. This last step is not always realised because of cost and time constraints.

14.5.4 Case studies

The materials used in earthworks and road construction have been the most studied until now, by methods and tools usually employed for the environmental assessment of wastes. There are very few leaching-scenario studies on materials used, for example, in house-type scenarios. There are no complete studies for human health risk assessment linked to the leaching of pollutants from construction materials in real utilisation scenarios. The existing studies concern mostly the leaching properties of materials and very few studies are extended to pollutant dispersion in soil and groundwater.

Globally, the research studies available in the literature concern different construction materials containing different wastes in various proportions and destined for particular utilisation scenarios. It is therefore not possible to generalise the results and conclusions, so several example studies are mentioned hereafter in their particular context.

Abbott, J., Coleman, P., Howlett, L., Wheeler, P., Environmental and Health Risks Associated with the Use of Processed Incinerator Bottom Ash in Road Construction. 2003, 2003 Report AEAT/ENV/R/0716. Available from www.breweb.org.uk (accessed 10 March 2006).

, ADEME. Assessment of the ecocompatibility of waste storage and recycling scenarios – a practical guide (in French, 2002. [Reference 4445, ADEME Editions, Angers, France.].

ADEME. State of the art of the use of non hazardous waste in cement plants/ Etat de l’art de la valorisation énergétique des déchets non dangereux en cimenteries. Rapport final, December. 2009; 2009:69.

Aubert, J.E., Husson, B., Sarramone, N. Utilization of municipal solid waste incineration (MSWI) fly ash in blended cement. Part 2. Mechanical strength of mortars and environmental impact. Journal of Hazardous Materials. 2007; 146:12–19.

Azizian, M.F., Nelson, P.O., Thayumanavan, P., Williamson, K.J. Environmental impact of highway construction and repair materials on surface and ground waters: Case study: crumb rubber asphalt concrete. Waste Management. 2003; 23:719–728.

Barna, R., Blanc, D. G2080v2,. Stabilisation-solidification des déchets, Techniques de l’ingénieur, 2011:17.

Barna, R., Bae, H.R., Méhu, J., van der Sloot, H., Moszkowicz, P., Desnoyers, C. Assessment of chemical sensitivity of Waelz slag. Waste Management. 2000; 20:115–124.

Barna, R., Rethy, Z., Imyim, A., Perrodin, Y., Moszkowicz, P., Tiruta-Barna, L. Environmental behaviour of a construction made of a mixture of hydraulic binders and air pollution control residues from municipal solid waste incineration. Part 1. Physico-chemical characterisation and modelling of the source term. Waste Management. 2000; 20:741–750.

Barna, R., Moszkowicz, P., Gervais, C. Leaching assessment of road materials containing primary lead and zinc slags. Waste Management. 2004; 24:945–955.

BRGM. Management of mining, quarrying and ore-processing waste in the European Union, 2001. [report BRGM/RP-50319-Fr.].

Bröns-Laot, G., Barna, R., Tiruta-Barna, L., Méhu, J., Moszkowicz, P., Magnié, M.-C., Innovative mortars containing MSWI bottom ashes for quarry backfilling: environmental assessment of long term behaviourGaballah I., ed. Global Symposium on Recycling, Waste Treatement and Clean Technology, 2004. [ISBN 84-95520-04-4, Madrid, 745–755.].

pr EN 14735 CEN. Characterization of waste – preparation of waste samples for ecotoxicity tests, 2003:42. [European Committee for Standardization, Brussels,].

16098 CEN/TR. Construction products: Assessment of release of dangerous substances – Concept of horizontal testing procedures in support of requirements under the CPD, 2010.

Chan, B.K.C., Bouzalakos, S., Dudeney, A.W.L. Cemented products containing waste from mineral processing and bioleaching. Minerals Engineering. 2009; 22:1326–1333.

De Windt, L., Dabo, D., Lidelöw, S., Badreddine, R., Lagerkvist, A. MSWI bottom ash used as basement at two pilot-scale roads: Comparison of leachate chemistry and reactive transport modelling. Waste Management. 2011; 31:267–280.

Duchesne, J., Bérubé, M.A. Evaluation of the validity of the pore solution expression method from hardened cement pastes and mortars. Cement and Concrete Research. 1994; 24:456–462.

Commission Decision 2001/118/EC of 16 January 2001 amending Decision 2000/532/EC as regards the list of wastes EC. Official Journal of the European Communities L Legis, 2001. [47, 0001–31 (16 February 2001)].

EC. Communication from the Commission to the Council and the European Parliament on thematic strategy on the urban environment. C0M/2005/0718 final, Brussels, 11 January 2006. Available from http://ec.europa.eu/environment/index_en.htm, 2006. [(accessed 20 September 2011).].

EC/DG ENV. Final Report Task 2 -Management of C&D Waste. February 2011. Available from http://ec.europa.eu/environment/waste/construction_demo-lition.htm, 2011. [(accessed 11 July 2011).].

EC-EDG. European Union waste policy, LIFE Focus: A cleaner, greener Europe, 2004. [2004, 3–6, ISBN 92-894-6018-0, ISSN 1725–5619, 0ffice for 0fficial Publications of the European Communities, Luxembourg.].

(European Coal Combustion Products Association e.V.) ECOBA. available from http://www.ecoba.com, 2011. [(accessed 11 July].

Eikelboom, R.T., Ruwiel, E., Goumans, J.J.J.M. The building materials decree: an example of a Dutch regulation based on the potential impact of materials on the environment. Waste Management. 2001; 21:295–302.

EPA. Catalogue of Standard Toxicity Tests for Ecological Risk Assessment, ECO Update. Publication 9345.0-05I, Intermittent Bulletin March 1994, Vol 2, No 2. Available from http://www.epa.gov/oswer/riskassessment/ecoup/pdf/v2no2.pdf, 1994. [(accessed 20 September 2011).].

EPA. Guidelines for ecological risk assessment, EPA/630/R-95/002F, April 1998. Available from http://www.epa.gov/risk/guidance.htm, 1998. [(accessed 11 July 2011)].

EPA. Risk assessment. Available from http://www.epa.gov/risk/guidance.htm, 2011. [(accessed 11 July 2011)].

(European Topic Centre on Sustainable Consumption and Production) ETC/SCP. EU as a Recycling Society. Present recycling levels of municipal waste and construction and demolition waste in the EU. working paper 2/2009. Available from http://scp.eionet.europa.eu, 2009. [(accessed 11 July 2011).].

EU, Official Journal of the European Union Regulation (EU) No. 305/2011 of the European Parliament and of the Council of 9 March 2011 laying down harmonized conditions for the marketing of construction products and repealing Council Directive 89/106/EEC, 2011. [L 88, 5–43 (4 April 2011).].

European Council Decision, concerning the European Catalogue of Hazardous Wastes 94/904/EC of 22 December, 1994.

EUROSLAG (The European Slag Association). Legal Status of Slags. Position Paper. January 2006. Available from http://www.euroslag.org, 2006. [(accessed 11 July 2011).].

François, D. An approach toward an assessment methodology of alternative materials for road construction, In 1st International Conference on Engineering for Waste Treatment. Beneficial Use of Waste and By-products. France: Albi; 2005.

François, D., Pierson, K. Environmental assessment of a road site built with MSWI residue. Science of the Total Environment. 2009; 407:5949–5960.

French Ministry of Environmen. Directorate for Prevention Pollution and Risk Control/Products and Division of Wastes, 1998. [Proposal for a Criterion and Evaluation Methods of Waste Ecotoxicity. Paris.].

GTZ-Holcim Public Private Partnership. Guidelines on Co-processing Waste Materials in Cement Production, 2006. [Copyright © 2006 Holcim Group Support Ltd and Deutsche Gesellschaft für Technische Zusammenarbeit (GTZ) GmbH. Available from http://www.holcim.com/ (accessed 1 September 2011).].

Hartlen, J., Fällman, A.M., Back, P.E., Jones, C. Principles for risk assessment of secondary materials in civil engineering work. Stockholm: Swedish EPA; 1999.

Hjelmar, O., van der Sloot, H.A., Gouyonnet, D., Rietra, R.P.J.J., Brun, A., Hall, D., Development of acceptance criteria for landfilling of waste: An approach based on impact modelling and scenario calculationsChristensen T.H., Cossu R., Stegmann R., eds. Sardinia 2001, Proceedings of the Eighth International Waste Management and Landfill Symposium, S. Margharita di Pula, 2001:712–721. [Cagliari, CISA,Vol. III,].

ISO. ISO standards/ by ICS/ 13. Environment, Health Protection, Safety. Available from http://www.iso.org/iso/iso_catalogue/catalogue_ics/catalogue_ics_browse.htm?, 2011. [ICS1 = 13 (accessed 20 September 2011).].

Jeong, S.-M., Osako, M., Kim, Y.-J. Utilizing a database to interpret leaching characteristics of lead from bottom ashes of municipal solid waste incinerators. Waste Management. 2005; 25:694–701.

Lapa, N., Barbosa, R., Morais, J., Mendes, B., Méhu, J., Santos Oliveira, J.F. Ecotoxicological assessment of leachates from MSWI bottom ashes. Waste Management. 2002; 22:583–593.

Marion, A.-M., De Laneve, M., De Grauw, A. Study of the leaching behaviour of paving concretes: Quantification of heavy metal content in leachates issued from tank test using demineralized water. Cement and Concrete Research. 2005; 35:951–957.

OECD. Guidelines for the Testing of Chemicals. Available from http://www.oecd-ilibrary.org/content/package/chem_guide_pkg-en, 2011. [(accessed 20 September 2011).].

Ore, S., Todorovic, J., Ecke, H., Grennberg, K., Lidelöw, S., Lagerkvist, A. Toxicity of leachate from bottom ash in a road construction. Waste Management. 2007; 27:1626–1637.

Pagotto, C., Béchet, B., Lanini, S., Descat, M., Paris, B., Piantone, P., Raimbault, G., Environmental impact assessment of the use of municipal solid waste incineration bottom ash in roadworkWASCON 2003, ed.. Waste Materials in Construction. San Sebastian, Spain, 2003.

Petkovic, G., Engelsen, C.J., Haoyad, A.O., Breedveld, G. Environmental impact from the use of recycled materials in road construction: method for decision-making in Norway. Resources Conservation and Recycling. 2004; 42:249–264.

Rakotoarisoa, Z., Prediction of the environmental behaviour of thermal process residues used as construction materials (in French), 2003. [PhD thesis, INSA Lyon, 252 pp.].

Reid, J.M., Evans, R., Holnsteiner, R., Wimmer, B., Gaggl, W., Berg, F., Pihl, K.A., Milvang-Jensen, O., Hjelmar, O., Rathmayer, H., François, D., Raimbault, G., Johansson, H.G., Hakansson, K., Nilsson, U., Hugener, M. Alternative materials in road construction. Final Report for the EU Commission. UK: Crowthorne; 2001.

Römbke, J., Moser, T., Moser, H. Ecotoxicological characterisation of 12 incineration ashes using 6 laboratory tests. Waste Management. 2009; 29:2475–2482.

Schiopu, N., Caractérisation des émissions dans l’eau des produits de construction pendant leur vie en oeuvre, 2007:278. [PhD thesis, INSA Lyon,].

Schiopu, N., Jayr, E., Méhu, J., Barna, L., Moszkowicz, P. Horizontal environmental assessment of building products in relation to the construction products directive (CPD). Waste Management. 2007; 27:1436–1443.

Schiopu, N., Tiruta-Barna, L., Jayr, E., Méhu, J., Moszkowicz, P. Modelling and simulation of concrete leaching under outdoor exposure conditions. Science of the Total Environment. 2009; 407:1613–1630.

Spanka, G., Thielen, G. Untersuchung zum Nachweis von verflussigenden Betonzusatemitteln und zu deren Sorptions-und Elutionsverhalten. Beton. 1995; 45:320–327.

Tiruta-Barna, L., Barna, R., Moszkowicz, P. Modelling of solid/liquid/gas mass transfer for environmental evaluation of cement-based solidified waste. Environmental Science and Technology. 2001; 35:149–156.

Tiruta-Barna, L., Rakotoarisoa, Z., Méhu, J. Assessment of the multi-scale leaching behaviour of compacted coal fly ash. Journal of Hazardous Materials. 2006; 137:1466–1478.

van der Sloot, H.A. Characterization of the leaching behaviour of concrete mortars and of cement-stabilized wastes with different waste loading for long term environmental assessment. Waste Management. 2002; 22:181–186.

Wilke, B.-M., Riepert, F., Kocha, C., Kuhne, T. Ecotoxicological characterization of hazardous wastes. Ecotoxicology and Environmental Safety. 2008; 70:283–293.