Solidification of contaminated sediments using GGBS-based binders

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Projects for DM in France and Ireland

The most important dredging sites in France are Port of Nantes St-Nazaire (10 MM.m3), Port of Bordeaux (7 MM.m3), Port of Rouen (5 MM.m3), Port of Le Havre (1,5 MM.m3), Port of Dunkirk, and Port of Marseille. Different types of inland management of DM in France were carried out such as land improvement, filling material for forestry, landfill capping, and beach nourishment.
(Harrington et al., 2013) reported the different types of applications and practices performed to date in their guidance on the beneficial use of sediments in Ireland. Among them are beach nourishment projects, landfill cover (e.g. Dublin Royal Canal Dredging Project), coastal protection (e.g. breakwater constructions in Ireland), concrete manufacture (use as a raw material as aggregates for concrete manufacture, Caladh Mor), etc.

Hierarchy of dredged materials

(Apitz, 2010) in her article discusses the waste hierarchy of dredged materials (Fig. I.3) as an important factor to help decision makers adopt the most relevant waste treatment technology for sustainable management. She proposes the following definitions for the main classifications of the dredged materials hierarchy:
! Pollution Prevention: different strategies can be examined to prevent pollution in harbors – the reduction or refusal of dredging procedures as well as controlling contaminant sources. Some approaches may help to avoid or reduce the dredging footprint, for example by using 2D or 3D site surveys in order to represent a more detailed contaminant distribution, to carry out more accurate and rigorous dredging procedures, or to reduce the volume of dredged material.
! Re-use : dredged material can be re-used depending on its contaminant concentration, therefore depending on the local or other level regulations. One of the main criteria is that sediments must stay in the same form as before treatment (e.g. sorting, cleaning). Dredged sediments may be relocated to maintain sediment balance in the environment without any risks, so this dredged material will be classified as re-used.
! Recycling : this strategy describes the use of dredged sediments in other forms than the original state – for example it can be used as a raw material for aggregate production, brick manufacturing, ceramics, etc. Recycling strategies are usually more expensive than disposal strategies, however contaminated sediments can present an attractive approach for construction materials production due to transport, energy, and resources economies.
! Recovery : the recovery strategy describes the beneficial use of contaminated or clean waste material if “biomass or energy is recovered” (e.g. materials that can be used as fuels).
! Disposal : this option is on the bottom of the waste hirarchy pyramid and is considered as the last resort for contaminated dredged material. Disposal of sediments may require additional monitoring due to possible release of contaminants over time as well as may be restricted by additional space needs. Among categories listed by (Apitz, 2010) are uncontrolled marine diposal (not for contaminated sediments), confined disposal facilities, disposal at sea with capping with clean sediment, disposal in special geotextile bags or in impermeable basins.

Origin and composition of sediments

Geological origin of sediments

The composition of sediments is highly dependent on the geology of the basin, topography, climate, and vegetation of a region. It is a highly dynamic and heterogeneous system (Schulz & Zabel, 2006). Regarding the geological provenance of sediments and their constituents’ formation, there are four main groups that can be distinguished:
:! Lithogenous (terrigenous) sediments that come from crushed and dissolved continental rocks (through chemical and physical weathering taking into account climatic changes, biological activity, etc.); the geographical location (relief, surface area, land-use) plays a considerable role in the particles transport and disposal; eolian transport plays a non-negligible role in the sediments’ transport
– the wind transports the fine fraction of sediments – clays and silts;
:! Biogenous sediment – the part of sediment that consists mostly of calcareous, siliceous, or phosphatic minerals which were formed in the biosphere. Some compounds of iron, aluminium, manganese, calcite, Mg-Calcite, or aragonite were formed by different groups of marine organisms (plankton, benthos).
:! Hydrogenous (Authigenic) sediment – represents the new formations by precipitation, alteration of particles in solution or within the sediment;
:! Anthropogenic sediments are formed due to human activity – port activities and urban areas produce industrial (mining industry, chemical industry, construction etc.), agricultural wastes containing the organic matter, pollutants from wastewater and sewage treatment.

Composition of sediments

A knowledge of a sediment’s chemistry, meaning its organic and inorganic constituents, gives an understanding of the chemical reactions that may take place from the point of view of contamination of the environment. These reactions may alter the solubility, mobility, and bioavailability of pollutants in waters (Sparks, 2002).
A sediment’s matrix consists of three main phases (Fig. I.4):
– Inorganic phase,
– Organic phase,
– Liquid phase.

Inorganic phase

The inorganic constituents of sediments possess different physical and chemical properties and can differ significantly in size – from clay’s fine fraction (<2μm) to gravel (>2 mm) and rocks. (Sparks, 2002) distinguished primary and secondary minerals based on their formation – primary minerals were not transformed chemically since their deposition and crystallization from the liquid state (lava); these minerals are sands (particle diameters between 0.05 and 2 mm) and silts (particle diameters between 0.002 and 0.05 mm). When the primary mineral’s structure was affected by weathering (by dissolution), secondary minerals were formed. These minerals are commonly called aluminosilicates. They are clay minerals such as kaolinite, montmorillonite, oxides such as gibbsite, sulfurs, carbonate minerals, etc.


A large amount of chemical reactions are impacted by sediments’ secondary minerals. Clay minerals present a complex structure of tetrahedral and octahedral sheets. The linkage of one octahedral and one tetrahedral sheet forms a 1:1 clay mineral with an ideal formula of Si4IVAl4VIO10(OH)8. A 2:1 clay mineral presents the structure when two tetrahedral sheets are coordinated to one octahedral sheet – the formula is Si8IVAl4VIO20(OH)4. In the interlayer space there are individual cations or cations octahedrally bound with hydroxyls (e.g. chlorites).
An isomorphous substitution phenomenon may take place during the formation of clay minerals. Depending on the cationic radius, different substitutions may occur. For example, in the octahedral sheet Fe2+, Fe3+, Mg2+, Ni2+, Zn2+, or Cu2+ can substitute for Al3+; or in the tetrahedral sheet Al3+ can substitute for Si4+ developing the charge imbalance compensated by cations (Sparks, 2002).
Figure I.5 shows the main structural schemes of secondary minerals. Here are some well-defined clay minerals and their properties:
1:1 Clay Ex. Kaolinite. Its structure has a silica tetrahedral sheet bonded to an aluminium octahedral sheet assembled by hydrogen bonding; the ideal chemical formula is Si4IVAl4VIO10(OH)8; no interlayer bonding is possible for a 1:1 clay mineral.
2:1 Clay Ex. Montmorillonite. The cations for the tetrahedral sheet are Si4+ and for the octahedral sheet the cations are Al3+, Fe2+, and Mg2+ with a chemical formula M0.33, H2OAl1.67(Fe2+, Mg2+)0.33 Si4O10(OH)2 where M indicates a metal cation in the interlayer space between sheets, being either Na+, Ca2+ or Mg2+ as the dominant cations. Montmorillonite clay is characterized by the presence of a big amount of water molecules between the sheets. This makes this type of clay mineral sensitive to swelling or shrinkage. Montmorillonite presents a high cation exchange capacity and high specific surface.
2:1 Clay Ex. Illite. The chemical characteristics of illite are close to mica minerals and smectite minerals; one-fourth of the tetrahedral atoms are Al3+ and illite has more Si4+ and Mg2+ than muscovite. Therefore, the negative charge of the isomorphous substitution is balanced by the K+ cations in the interlayer space, as well as by Ca2+, Mg2+, or NH4+, but less often. This type of clay is non-expanding and has a low CEC capacity (Sondi et al., 1996).

Oxides, hydroxides, oxyhydroxides

Iron, aluminium, and manganese oxides play a crucial role in the chemistry of sediments regarding heavy metal mobility. Gibbsite (Al(OH)3) and boehmite (γ-AlOOH) are the most common Al-oxides naturally present in soils and sediments. For Fe oxides, goethite (FeO(OH)) is one of the most prevalent and thermodynamically stable minerals and takes the form of needle-shaped crystals. The other common Fe oxide is hematite. Some heavy metal cations can be found in Fe oxides – Ni, Ti, Co, Cu, Zn or Fe can be isomorphically substituted by Al, Mn, and Cr (Sparks, 2002).


Carbonate minerals are much more soluble compared to siliceous minerals. The most often encountered carbonates in soils are calcite (CaCO3) and magnesite (MgCO3).
Due to their instability, they can be converted to dolomite (CaMg(CO3)2), ankerite ((Ca,Fe,Mg)2(CO3)2) or siderite (FeCO3) (Sparks, 2002).
The solubility of carbonates partially controls the pH of the sediment’s solution due to the buffer effect. Therefore, at some concentration of carbonates, the pH of the sediment can be slightly alkaline (∼8) and some heavy metals may precipitate in the form of carbonates (hydroxycarbonates) – for example the zinc cation may precipitate in the form of ZnCO3 or Zn5(OH)6(CO3)2 species and copper in the form of malachite (Cu2(OH)2CO3) or azurite (Cu3(OH)2(CO3)2) (Cazalet, 2012; Kribi, 2005).


Sulfurs in sediments originate from the reduction process of sulfates ions in an anaerobic medium (Couvidat, 2016). The reduction reactions of sulfates in the sediment matrix are controlled biologically through the action of sulfate-reducing bacteria in the presence of reactive organic matter and without oxygen, (1) (Cazalet, 2012; Lesven, 2008):
The formation of sulfurs plays a key role in the mobility of some trace elements in sediments like Fe, Mn, Pb, Cd, Hg, Cu, As, and Zn. Depending on the chemical and microbiological environment within the sediment matrix, the different oxidation states for sulfur may occur – from sulfates (+VI) in the oxygenated water-sediment interface to sulfurs in the most reduced form (-II). According to (Lesven, 2008) the principal dissolved species of sulfur is HS- in the natural environment. Subsequently, there are several reactions than can take place – the formation of organo-sulfur compounds, the precipitation of iron sulfide (pyrite) or the complexation of heavy metals, (2), (3) (Billon, 2001):
During the dredging and dewatering process, inorganic sulfurs may undergo oxidation and further sulfate formation. For example in the case of the reaction of pyrite with oxygen, the pH decreases due to the formation of sulfuric acid (4), (Gamsonré, 2014):
The formation of sulfuric acid can produce the reaction of this acid with carbonates.
Consequently, gypsum formation occurs following reaction (5) (Gamsonré, 2014):
The further crystallization of gypsum may become an important issue for sediment recycling and the risks of swelling and expansion should be taken into account. In the case of the S/S method, an appropriate binder should be designed in order to avoid degradation over time due to volumetric changes.

Organic phase

The organic compounds of marine geochemistry present a complex environmental system that controls different important processes in sediments. The large structural variety of the organic compounds turns out to be an important challenge to be quantified and distinguished as different groups and components co-exist. Most of the molecules in organic matter (OM) form macromolecules, for example proteins and polysaccharides, that have to be disassembled into smaller structural units – amino acids and sugars (Emerson & Hedges, 2008).
The organic structure is built up of carbon-linked substructures of single molecules, such as hydrocarbons that have only carbon chains with hydrogen adjuncts and are considered as one of the simplest organic molecules.
Organic matter is divided into two main groups of substances: humic and non-humic. Non-humic material contains carbohydrates, proteins, peptides, amino acids, fats, waxes, and low-molecular-weight acids. These OM components are not very stable and can be easily transformed by microorganisms (Sparks, 2002). Humic substances (HS) are subdivided into humic acid (HA), fulvic acid (FA), and humin, which have different solubilities in different pH mediums. Fulvic acid and humic acid are soluble in alkaline media (characteristic of hydraulic binders) but not humin (Fig. I.6). Humic substances vary considerably in molecular weight as well as in size. According to (Piccolo, 1996; Stevenson, 1982) HS present as two- or three-dimensional macromolecules (Fig. I.7) that are interconnected and form a negatively charged surface from the ionization of acidic functional groups, for example carboxyls.
In geochemistry the “biomarkers” method is used to determine the properties and provenance of natural organic systems. Regarding organic carbon compounds, the main biomarkers are proteins, lipids and carbohydrates in living organisms, and other biopolymers such as lignin and tannin from plants (Emerson & Hedges, 2008). Concerning the protein amino acids group, these compounds are some of the most abundant in marine biomass, having carboxyl (COOH), amino (CNH2), hydrogen, and R groups. In the range of neutral pH, the amino acid group will present the ionized forms of COO- and CNH3+. The next most abundant biomarkers are carbohydrates with the general formula (CH2O)n. Carbon sugars are monomeric units of carbohydrates, combining in oligo- or polysaccharides. The next group of biomarkers consists of lipids, assembles hydrocarbons, fatty acids, sterols, and alkenones families. Lignins and tannins form the next group of biomarkers which represent the phenol biopolymer as products of vascular plants.
One of the types of marine organic matter is dissolved organic matter (DOM) which is a form of dissolved carbon in water. This phase contains mostly biopolymers and can be chemically or biologically degraded in the water column (Cazalet, 2012). As DOM presents a source of bioactive elements, it plays an important role in photochemical reactions, metals complexation, etc.

Contaminants in sediments and their sources

Toxic elements are widespread within coastal and estuarine sediments and make up part of the complex environmental chemistry of sediments nowadays. Heavy metals, polyaromatic hydrocarbons, organochlorines (such as PCBs), pesticides, etc. may present a considerable risk for aquatic life and human health. Due to industrialization processes developed through the last decades, the heavy metals level in sediments is more elevated than naturally occurring metals in sediment minerals.
Trace elements (<0.1% in natural materials) are considered as toxic if they exceed concentrations presenting risks for living organisms. They include trace and heavy metals, metalloids, micronutrients, and organic contaminants (Sparks, 2002). The source of these elements can originate from natural material as well as from anthropogenic activities – industrial discharges, mining activities, pesticides, agriculture, and harbor activities (e.g. boat painting).
Regarding the complexity and heterogeneity of sediment composition and chemistry, there is evidence of their strong impact on the bioavailability and fate of toxic elements in the environment. It can be found in the literature that the term “sink” is widely used to emphasize the great capacity of sediments to retain and stock metals. Pollutant mobility and bioavailability are strongly dependent on their chemical speciation. In this part of the literature review, some main pollutants and their chemical and mineral forms occurring in sediments will be considered. Unlike organic molecules, the main problem with inorganic contaminants is that they cannot be biodegraded, therefore heavy metals have a tendency to bioaccumulate (Couvidat, 2016). In this part of the review, some toxic elements encountered in Dublin’s sediment matrix are considered, specifically their sources and behaviour. The pe-pH (Pourbaix) diagrams were calculated at standard ambient temperature and pressure (25°C; 101.325 kPa).

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Organic contaminants

The environmental risks related to organic contaminants increased sharply through the last decades. Anthropological activities generate a high rate of organic pollutants from different sources: industrial leaks and spills, improper application of pesticides, leaks from oil and chemical storage tanks, leaks from pipelines, accidents, spills during transportation, etc. (Snousy, 2017). The primary source of groundwater contamination is petroleum hydrocarbons from underground storage tanks. Different physical forms of organic contaminants are presented in Fig. I.8. Most organic molecules will be adsorbed to the soils and sediments surface due to their hydrophobic nature.
In the aquatic medium, the most widespread organic contaminates are PAHs (polycyclic aromatic hydrocarbons), PCBs (polychlorinated biphenyls) and different organometallic compounds such as TBT (tributyltin chloride) (Fig. I.9). TBT compounds were accumulated in sediments due to their large use as anti-foulants until they were banned due to their high toxicity. They can induce disruption in the reproductive function in mammals, behave as hepatoxins, immunotoxins, neurotoxins, and obesogens (Haydee & Dalma, 2017).

Inorganic contaminants


Zinc is considered as an essential element for many metalloenzymes and it can be very toxic for plants at high concentrations. Zinc is found in sewage sludge at high concentrations. Its provenance is often associated with industrial waste and metal factories. In harbor areas, zinc is applied on the submerged parts of boats as an antifouling agent/paint (Alzieu, 1999).
Regarding the mineralogy of zinc in sediments, it is often combined with lead or copper, as well as cadmium (Burnol et al., 2006; Lesven, 2008). It can be readily adsorbed onto clay minerals, carbonites, or hydroxides which retards its desorption into the environment (Burnol et al., 2006). (Tessier, 1979) reported an important affinity of zinc to iron and manganese oxides according to the following order: FeO/MnOOH>carbonates> clays.
Zinc is found in nature in the oxidation state of +2. The most abundant mineral forms of zinc are zincite (ZnO), zinc carbonate (smithsonite, ZnCO3), sphalerite, or zinc sulfide (ZnS), as well as silicates and mixed oxides of zinc and iron. Zinc sulfides are the main insoluble form of zinc precipitate that is formed in anaerobic conditions (Burnol et al., 2006). The different ionic species of zinc that can be found in soils are presented in Fig. I.10 (Burnol et al., 2006).


Different types of nickel minerals are found in the geochemistry of soils and sediments. They can be found in the form of oxides, carbonates, and silicates and are particularly abundant in the form of sulfides (vaesite (NiS2), millerite (NiS), and iron nickel sulfide (pentlandite, Fe,Ni9S8)) (Gamsonré, 2014). Nickel minerals are poorly soluble in natural systems. Nickel salts can also occur in the natural environment and nickel chlorides, sulfates, and carbonates are more soluble than nickel oxide. The presence of nickel in the aquatic environment can be attributed to the natural phenomena of volcanism or forest fires, as well as to sources from human activities such as the metal industry and petroleum combustion (Alzieu, 1999), (Lesven, 2008).
Nickel has an important affinity to iron oxides via adsorption (e.g. goethite), kaolinite, and organic matter (Gamsonré, 2014). However, it is readily exchangeable because of the attached hydrated nickel species on the surfaces of inorganic or organic phases, governed by the electrostatic forces through hydrogen bonding (Rinklebe & Shaheen, 2017).
In aquatic media, nickel is present only in the oxidation state of +2. In anoxic sediments nickel will precipitate with HS- to form nickel sulfides. In acidic or neutral conditions, the concentration of nickel is dependent on the solubility of solid carbonates (Fig. I.11).


Cadmium is abundantly present in the environment and is both highly toxic and carcinogenic. (Kubier et al., 2019) reported in his review that cadmium is one of the most mobile toxic elements. It can easily replace calcium in minerals and consequently human bodies due to the similar ionic size and chemical behaviour. The anthropogenic sources of cadmium are mining, the metal industry, the petroleum industry, textiles, etc. (Fig. I.12).
Cadmium in nature is often associated with zinc and can substitute for Zn in sphalerite (ZnS) or smithsonite (ZnCO3). Some other cations can also participate in isomorphous substitution with cadmium – Ca, Fe, Zn, and Pb (Kubier et al., 2019). In the anoxic sulfurs environment, cadmium can be present as the metal sulfide mineral greenockite (CdS). Adsorption is the main mechanism of cadmium immobilization/retention – it can be adsorbed onto the clay minerals, iron or manganese oxides, carbonates, or even associated with organic matter and sulfide complexes (Burnol et al., 2006). The most stable cadmium phase in reducible media is CdS.
When cadmium ions are present in oxidation conditions, the solubility will be controlled by cadmium carbonate (CdCO3) at neutral pH, by sulfates/hydroxides at slightly more alkaline pH, and cadmium hydroxides will predominate in highly alkaline media (Fig. I.13). In the aquatic environment, the free ion Cd2+ is the most abundant form of the hydrolysable cation.


According to the (OSPAR Commission, 2016), the main sources of copper in the environment are the mining industry (∼18 million tonnes per year), chemical production (chlorides, sulfides, and oxides), and electronics production. In the marine environment copper originates from boat propellers or chemicals used for antifouling treatments. Once TBT compounds were forbidden as efficient antifouling agents due to their high toxicity, copper use in boat paints and biocides increased significantly. The main copper minerals found in nature are chalcocite copper(I) sulfide (Cu2S) and copper iron sulfide (chalcopyrite, CuFeS2). Some oxide minerals such as cuprite (Cu2O), tenorite (CuO), and carbonate minerals such as malachite (Cu2CO3(OH)2) and azurite (Cu3(CO3)2(OH)2) can also be found (Cazalet, 2012).
Different sediment phases may adsorb copper in natural conditions – Cu ions have a great affinity for organic matter, iron and manganese oxides, clays, and organoclay complexes. In aquatic media the oxidation state of copper is +1 or +2 depending on the physicochemical conditions of the environment. When copper is in the oxidizing medium, the main phases controlling its availability are carbonites or oxides depending on the pH (Fig. I.15). Copper sulfides will precipitate in the reducing environment in the presence of sulfurs (Fig. I.14).
The toxicity of copper depends on its oxidation state and chemical form. According to (Alzieu, 1999) copper is more toxic in the oxidation state of +1. (Fairbrother et al., 2007) reported that for some marine species even 2µg/L of copper can be critical and, in some cases, lethal.


Arsenic can be found in nature in different oxidation states (-3, 0, +3 and +5), meanwhile in the aquatic environment arsenic occurs mostly as oxyanions of trivalent arsenite (As(III)) or pentavalent arsenate (As(V)). Some organic forms can also be found in sediments when industrial pollution is high. Inorganic arsenic naturally occurs in groundwater; it is used in the processing of glass, pigments, textiles, paper, metal adhesives, wood preservatives, and ammunition (Arsenic). Inorganic arsenic compounds are considered to be highly toxic and carcinogenic.
More than 200 As minerals have been identified – among them elemental arsenic, arsenides, sulfides, oxides, arsenates and arsenites (Smedley, 2005). Arsenic in minerals is often found to be associated with other metals and one of the most abundant is arsenopyrite (FeAsS). Arsenic mobilization is controlled by its association with oxide minerals – many studies were conducted on the adsorption of arsenite and arsenate onto hydrous ferric oxides, aluminium and manganese oxides, and adsorption on clays minerals due to their oxide-like character of edges, e.g. kaolinite and montmorillonite (Burnol et al., 2006). The stabilization of As is a highly pH dependent process.
The reduction of iron oxides produced in anaerobic sediments is an important phenomenon for arsenic mobility – adsorbed or combined arsenic with hydrous iron oxides will be dissolved.

Table of contents :

Chapter I: Literature review
I.1 Introduction
I.2 Management of harbor sediments
I.2.1 Defining dredging operations
I.2.2 European management of sediments. Legislation
I.2.3 Existing projects and hierarchy for DM management
I.2.3.1 Projects for DM in France and Ireland
I.2.3.2 Hierarchy of dredged materials
I.3 Origin and composition of sediments
I.3.1 Geological origin of sediments
I.3.2 Composition of sediments
I.3.2.1 Inorganic phase
I. Clays
I. Oxides, hydroxides, oxyhydroxides
I. Carbonates
I. Sulfurs
I.3.2.2 Organic phase
I.4 Contaminants in sediments and their sources
I.4.1 Organic contaminants
I.4.2 Inorganic contaminants
I.4.2.1 Zinc
I.4.2.2 Nickel
I.4.2.3 Cadmium
I.4.2.4 Copper
I.4.2.5 Arsenic
I.4.2.6 Chromium
I.5 Factors affecting HM availability
I.5.1 Influence of pH
I.5.2 Oxidation-Reduction Potential
I.5.3 Cation exchange capacity (CEC)
I.5.4 Organic matter
I.5.5 Salinity
I.6 Treatment technologies for dredged sediments
I.6.1 Pre-treatment
I.6.2 Physical separation
I.6.3 Washing
I.6.4 Electrokinetic remediation
I.6.5 In-situ Capping
I.6.6 Biological remediation
I.6.7 Thermal Extraction
I.6.8 Solidification/Stabilization (S/S)
I.6.8.1 Tests for the S/S evaluation
I.6.8.2 S/S main reactions mechanisms of HM fixation
I.6.8.3 Stability of the treated matrix
I.7 Considered binding agent properties
I.7.1 Ordinary Portland cement (OPC)
I.7.1.1 Composition of OPC
I.7.1.2 Hydration chemistry of OPC
I.7.2 Ground granulated blast-furnace slag (GGBS)
I.7.2.1 What is GGBS?
I.7.2.2 Chemical composition, mineralogy of GGBS
I.7.2.3 Hydration of GGBS-based binders
I.7.3 C-S-H structure
I.7.4 OPC and GGBS-based binders in S/S practice
I.7.4.1 OPC as a S/S agent
I.7.4.2 GGBS as a S/S agent
Chapter II: Materials and Methods
II.1 Introduction
II.2 Raw materials
II.2.1 Sediment
II.2.1.1 Origin of the considered sediments
II.2.1.2 Physical analysis of the sediments
II. Particle size distribution
II. Density
II.2.1.3 Mineralogy
II. XRD analysis
II. TGA (Thermogravimetric analysis), DTA (Differential thermal analysis)….
II.2.1.4 Chemical properties of the sediments
II. pH measurements
II. Total Organic Carbon (TOC) analysis
II. Cation Exchange Capacity (CEC) measurement
II.2.1.5 Operational fractionation of inorganic pollutants
II. Total Attack of the Dublin sediment and the main binding agents
II. Total attack procedure
II. Results of the total attack of the Dublin sediment
II. Enrichment factor
II. Sequential extraction of HM from the Dublin sediment
II. Sequential extraction fractions proposed by Tessier protocol
II. Procedure of sequential extraction applied in the study
II. Sequential Extraction results for the major elements
II.2.2 Binders and other materials
II.3 Sediment-Binder Systems
II.3.1 Samples’ preparation
II.3.1.1 Samples for the main case Dublin port sediment
II.3.1.2 Samples for the study of the impact of the nature of sediments
II.3.2 Samples’ characterization
II.3.2.1 Compressive Strength
II.3.2.2 Shrinkage Test
II.3.2.3 Leaching Test
II.3.2.4 XRD analysis
II.3.2.5 Zeta Potential
II.3.2.6 Rheology
II.3.2.7 Tomography
II.3.2.8 SEM analysis
II.3.2.9 Solid State Nuclear magnetic resonance
II.3.2.10 Pyrolysis GC/MS
II.3.2.11 XAS (X-ray Absorption Spectroscopy)
II.4 Simplified models
II.4.1 Clay-Binder system
II.4.1.1 Materials
II.4.1.2 Samples’ preparation
II.4.1.3 Samples’ characterization
II. Compressive strength
II. XRD analysis
II. 27Al MAS NMR investigation
II. SEM analysis
II. Zeta Potential
II. Rheology
II.4.2 Heavy metals – binders’ system for the investigation of early hydration
II.4.2.1 Materials
II.4.2.2 Samples’ preparation
II.4.2.3 Samples’ characterization
II. XRD analysis
II. 27Al and 29Si MAS NMR investigation
II. Isothermal calorimetry analysis
II. Zeta Potential measurements
II.4.3 Investigation of Ni and Cd in OPC and GGBS-based binders
II.4.3.1 Materials
II.4.3.2 Samples’ preparation
II.4.3.3 Samples’ characterization
II. Leaching test
II. X-ray Absorption Fine Structure (XAFS) of Ni and Cd in binders
Chapter III. Solidification of contaminated sediments using GGBS-based binders
III.1 Introduction
III.2 Main case – S/S treatment of the Dublin port sediment
III.2.1 Compressive strength evolution of the Dublin sediment mixed with binders at 150 kg/m3
III.2.2 pH of the mixtures
III.2.3 Mineralogical analysis
III.2.4 Shrinkage results
III.2.5 Microstructure investigation
III.2.5.1 X-Ray microtomography
III.2.5.2 SEM observations
III.2.6 Electrokinetic properties
III.2.6.1 Results
III.2.6.2 Discussion
III.2.7 Conclusions
III.3 Study of the various factors impacting the evolution of compressive strength during S/S treatment
III.3.1 Impact of the nature of sediments
III.3.1.1 Compressive strength results and pH measurements
III.3.1.2 XRD analysis of the Dublin and Gothenburg sediments mixed with the binders
III.3.1.3 Conclusions
III.3.2 Impact of the sediment pore water
III.3.3 Impact of the clay fraction. Clay-Binder system’s study
III.3.3.1 Compressive strength and pH of the clay-binder systems
III.3.3.2 XRD analysis of the clay-binder systems
III.3.3.3 27Al NMR analyses of the clay-binder systems
III.3.3.4 TGA
III.3.3.5 SEM observations
III.3.3.6 Impact of the binder’s ions on the rheology and zeta potential of the clay
III. Zeta Potential results of the clay-binder systems
III. Rheological measurements of the clay-binder systems
III.3.3.7 Discussion
III.3.4 Impact of dispersing agents. Dublin sediment – Binder system
III.3.4.1 Effect on the rheological behaviour
III.3.4.2 Compressive strength results
III.3.4.3 Zeta Potential measurements
III.3.4.4 Discussion
III.4 Interaction between the Dublin sediment’s organic matter and binders
III.4.1 Organic matter reconstruction before and after S/S treatment
III.4.2 13C – NMR investigation
III.4.3 Discussion
III.4.4 Effect of the organic matter on early age hydration of hydraulic binders
III.4.4.1 Heat flow results
III.5 Conclusions
Chapter IV: Impact of heavy metals pollutants contained in dredged sediments on the hydration of hydraulic binders
IV.1 Introduction
IV.2 Early age hydration of the binders with HM
IV.2.1 Impact of zinc
IV.2.1.1 100%OPC formulation
IV.2.1.2 85%GGBS/15%OPC formulation
IV.2.1.3 Supersulfated formulation
IV.2.1.4 Discussion
IV.2.2 Impact of copper
IV.2.2.1 100%OPC formulation
IV.2.2.2 85%GGBS/15%OPC formulation
IV.2.2.3 Supersulfated formulation
IV.2.2.4 Discussion
IV.2.3 Impact of nickel
IV.2.3.1 100%OPC formulation
IV.2.3.2 85%GGBS/15%OPC formulation
IV.2.3.3 Supersulfated formulation
IV.2.3.4 Discussion
IV.2.4 Impact of cadmium
IV.2.4.1 100%OPC formulation
IV.2.4.2 85%GGBS/15%OPC formulation
IV.2.4.3 Supersulfated formulation
IV.2.4.4 Discussion
IV.2.5 Zeta Potential of the binders in the presence of HM
IV.2.5.1 100%OPC formulation
IV.2.5.2 85%GGBS/15%OPC
IV.2.5.3 Discussion
IV.2.6 Conclusions
IV.3 27Al and 29Si NMR spectra of the ‘hydraulic binders-heavy metals’ system. Long term hydration
IV.3.1 Introduction
IV.3.2 Considered samples
IV.3.3 27Al NMR spectra of the OPC system in the presence of Cd and Zn
IV.3.4 27Al NMR spectra of the GGBS85 system in the presence of Cd and Zn
IV.3.5 27Al NMR spectra of the AAS system in the presence of Cd and Zn
IV.3.5.1 29Si NMR and 1H-29Si spectra of the AAS system in the presence of Zn
IV.3.5.2 Conclusions
Chapter V. Stabilization of heavy metals using GGBS-based hydraulic binders
V.1 Introduction
V.2 Batch leaching test
V.2.1 Leaching test Results
V.2.2 pH measurements
V.2.3 Impact of a strongly acidic and basic pH on the stability of HM in the Dublin sediment
V.2.4 Discussion
V.2.4.1 Oxyanions metals
V.2.4.2 Cationic heavy metals
V.3 XAS Investigations
V.3.1 Speciation of Zn and Cu in the raw Dublin sediment and the sediment mixed with the binders
V.3.1.1 Zinc speciation
V.3.1.2 Copper speciation
V.3.2 Speciation of Ni and Cd in the binders
V.3.2.1 Cadmium speciation
V.3.2.2 Cadmium Leaching
V.3.2.3 Nickel speciation and leaching results
V.4 Sequential Extractions of HM from the Dublin sediment with and without hydraulic binders’ addition
V.4.1 Untreated sediments samples
V.4.2 Impact of the addition of hydraulic binder on the fractionation of HM
V.4.3 Discussion
V.4.4 Conclusions
V.5 Conclusions


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