Environmental Soil Remediation and Rehabilitation: Existing and Innovative Solutions (Applied Environmental Science and Engineering for a Sustainable Future) 3030403475, 9783030403478

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Applied Environmental Science and Engineering for a Sustainable Future

Eric D. van Hullebusch  David Huguenot · Yoan Pechaud  Marie-Odile Simonnot  Stéfan Colombano  Editors

Environmental Soil Remediation and Rehabilitation Existing and Innovative Solutions

Applied Environmental Science and Engineering for a Sustainable Future

Series editors Jega V. Jegatheesan, School of Engineering, RMIT University, Melbourne, VIC, Australia Li Shu, LJS Environment, Melbourne, Australia Piet Lens, UNESCO-IHE Institute for Water Education, Delft, The Netherlands Chart Chiemchaisri, Kasetsart University, Bangkok, Thailand

Applied Environmental Science and Engineering for a Sustainable Future (AESE) series covers a variety of environmental issues and how they could be solved through innovations in science and engineering. Our societies thrive on the advancements in science and technology which pave the way for better standard of living. The adverse effect of such improvements is the deterioration of the environment. Thus, better catchment management in order to sustainably manage all types of resources (including water, minerals and others) is of paramount importance. Water and wastewater treatment and reuse, solid and hazardous waste management, industrial waste minimisation, soil restoration and agriculture as well as myriad of other topics needs better understanding and application. This book series aims at fulfilling such a task in coming years.

More information about this series at http://www.springer.com/series/13085

Eric D. van Hullebusch • David Huguenot • Yoan Pechaud • Marie-Odile Simonnot • Stéfan Colombano Editors

Environmental Soil Remediation and Rehabilitation Existing and Innovative Solutions

Editors Eric D. van Hullebusch Institut de physique du globe de Paris Université de Paris Paris, France Yoan Pechaud Laboratoire Géomatériaux et Environnement Université Gustave Eiffel Champs-sur-Marne, France

David Huguenot Laboratoire Géomatériaux et Environnement Université Gustave Eiffel Champs-sur-Marne, France Marie-Odile Simonnot Laboratoire Réactions et Génie des Procédés Université de Lorraine Nancy, France

Stéfan Colombano BRGM - Bureau de Recherches Géologiques et Minières Orléans, France

ISSN 2570-2165 ISSN 2570-2173 (electronic) Applied Environmental Science and Engineering for a Sustainable Future ISBN 978-3-030-40347-8 ISBN 978-3-030-40348-5 (eBook) https://doi.org/10.1007/978-3-030-40348-5 © Springer Nature Switzerland AG 2020 This work is subject to copyright. All rights are reserved by the Publisher, whether the whole or part of the material is concerned, specifically the rights of translation, reprinting, reuse of illustrations, recitation, broadcasting, reproduction on microfilms or in any other physical way, and transmission or information storage and retrieval, electronic adaptation, computer software, or by similar or dissimilar methodology now known or hereafter developed. The use of general descriptive names, registered names, trademarks, service marks, etc. in this publication does not imply, even in the absence of a specific statement, that such names are exempt from the relevant protective laws and regulations and therefore free for general use. The publisher, the authors, and the editors are safe to assume that the advice and information in this book are believed to be true and accurate at the date of publication. Neither the publisher nor the authors or the editors give a warranty, expressed or implied, with respect to the material contained herein or for any errors or omissions that may have been made. The publisher remains neutral with regard to jurisdictional claims in published maps and institutional affiliations. This Springer imprint is published by the registered company Springer Nature Switzerland AG. The registered company address is: Gewerbestrasse 11, 6330 Cham, Switzerland

Preface

Soils and groundwater may be contaminated by a large number of hazardous pollutants (both organic and inorganic compounds). The toxicity and the persisting characters of most of these pollutants lead to consider them as serious threats toward human health and the environment, requiring the application of remediation measures. Many established techniques are implemented to clean up, eliminate, recover, or sequester these hazardous pollutants from contaminated sites. However, even if many techniques are available and used at field scale, still, improved remediation approaches are currently being developed and tested on-site. The objectives are the reduction of the treatment costs as well as the improvement of remediation efficiencies. In this frame, this book provides a review on current research and development approaches dedicated to contaminated site remediation technologies. The newly developed remediation technologies for contaminated sites include bioremediation, physical/chemical remediation, and integrated remediation processes. The fundamentals and applied aspects of different newly developed remediation approaches are discussed and detailed in this book. Several case studies are also reported in different chapters (i.e., Chaps. 1, 2, 3, and 6). Chapter 1 provides the basic principles about the distribution of contaminants between the soil compartments and the transfer kinetics between these pools. Extracting agents (EAs), technologies, and methods used to mobilize priority contaminants are also reviewed by considering the performance, benefits, costs, and risks when implemented on-site. Soil washing (SW) enhanced by the use of EAs is seen as a promising technique for the removal of hydrophobic organic compounds such as polycyclic aromatic hydrocarbons, petroleum hydrocarbons, chlorinated solvents, and polychlorobiphenyls from contaminated soils. Other contaminants such as organic liquids (i.e., nonaqueous phase liquids, NAPLs) with low water solubility described as light NAPLs (i.e., LNAPLs) and dense NAPLs (i.e., DNAPLs) are very often encountered on contaminated sites. Removing free products is of primary importance for the remediation of such sites. Conventional NAPL remediation consists of pumping the free product until residual saturation is reached (i.e., no more NAPLs can be recovered). Decreasing this v

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residual saturation (i.e., increasing the recovery yield of the free product) may contribute to reducing (1) contaminants dissolved in water, (2) the duration of the remediation operation, (3) the extent of plumes and related contaminant concentration levels, and consequently, (4) remediation costs. Chapter 2 reports on conventional technologies (pump-and-treat, skimming, bioslurping, and recovery trenches) for free product recovery from the practical and theoretical viewpoints. Chapter 2 also describes the advantages and limitations of these techniques and discusses innovative approaches such as thermal and chemical enhancements (i.e., surfactants) aiming to increase free product recovery yields and rates. Thermal enhancements are being used for the remediation of contaminated sites for more than twenty years. Depending on the heating temperature, in situ thermal treatment may be used in combination with physical, chemical, and biological treatment processes. Chapter 3 describes how increasing the temperature contributes to (1) increasing the organic contaminant vapor pressure, the aqueous solubility, the Henry’s law constant, and the rate of (bio)chemical degradation and (2) decreasing the organic carbon partitioning, liquid density, liquid viscosity, and interfacial tension. Chapter 3 also examines the conventional technologies for in situ thermal enhancement from the theoretical and practical viewpoints including steam-enhanced extraction, thermal conductive heating, and electrical resistance heating. The advantages and limitations of these technologies, as well as innovative techniques such as radio frequency heating, are also discussed. Different treatment techniques applied to treat source zones contaminated with NAPLs are usually implemented. The most commonly implemented techniques are oxidation, sparging, surfactant flushing, and low-temperature thermal treatment. Chapter 4 discusses the efficiency of such techniques in the same experimental solid matrix and under the same experimental conditions at three different scales: batch, columns, and metric pilot tank experiments. Thermal treatment was reported as the most efficient remediation technique. However, a need to perform experiments at three experimental scales has been evidenced: batch studies allowing appropriate variations in experimental conditions, the column experiments permitting the optimization of surfactant and oxidant injection strategies, and tank experiments allowing us to investigate heterogeneous flow conditions. The use of organic amendments (e.g., digestates) as nutrient source for the bioremediation of petroleum hydrocarbon-contaminated soils is discussed in Chap. 5. This chapter provides in particular a review on the use of amendment for soil composting treatments and reports on the best monitoring approaches including chemical and biological assays as well as the use of molecular markers. The use of zero-valent iron for the degradation of chlorinated organic compounds (COCs) in soil and groundwater is described in Chap. 6. Even if chemical oxidation was first developed for in situ application, chemical reduction is currently one of the most important emerging remediation techniques for COC treatment. Chapter 6 describes the latest developments of in situ chemical reduction technologies aiming to enhance remediation rates. The influence of environmental conditions for in situ applications is reported and a case study is presented. Chapter 7 describes how manganese (Mn) oxides can be used for soil remediation. Based on the mineralogy

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as well as the reactivity of Mn oxides, these latter can be used as (1) oxidative reagents for organic and inorganic substances, (2) sorption mediums of trace metals and metalloids, and (3) reactive sorbent agents for chemical warfare agents and organophosphate pesticides. To conclude, we would like to convey our appreciation to all the contributors as well as the numerous reviewers who kindly accepted to provide comments for every chapter. Our special thanks to Mrs Judith Terpos, Mrs Alexandrine Cheronet, and Mrs Eva Loerinczi from Springer Nature for their kind support and great efforts in bringing the book to completion. We are glad to submit this book, and we hope that the readers will appreciate reading this volume as much as we enjoy working on this topic for more than 10 years. Paris, France Orléans, France Nancy, France Champs-sur-Marne, France Champs-sur-Marne, France November 20, 2019

Eric D. van Hullebusch Stéfan Colombano Marie-Odile Simonnot Yoan Pechaud David Huguenot

Contents

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Contaminant Mobilization from Polluted Soils: Behavior and Reuse of Leaching Solutions . . . . . . . . . . . . . . . . . . . . . . . . . . . . Nicolas Fatin-Rouge Free Product Recovery of Non-aqueous Phase Liquids in Contaminated Sites: Theory and Case Studies . . . . . . . . . . . . . . . Stéfan Colombano, Hossein Davarzani, Eric D. van Hullebusch, Ioannis Ignatiadis, David Huguenot, Sagyn Omirbekov, and Dominique Guyonnet

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In Situ Thermal Treatments and Enhancements: Theory and Case Study . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 149 Stéfan Colombano, Hossein Davarzani, Eric D. van Hullebusch, Ioannis Ignatiadis, Huguen Huguenot, Clément Zornig, and Dominique Guyonnet

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Comparing the Efficiency of Oxidation, Sparging, Surfactant Flushing, and Thermal Treatment at Different Scales (Batch, Column, Metric Pilot) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 211 Florie Jousse, Patrick Höhener, Grégory Cohen, and Olivier Atteia

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Potential Use of Waste-to-Bioenergy By-Products in Bioremediation of Total Petroleum Hydrocarbons (TPH)-Contaminated Soils . . . . . . 239 Anna Gielnik, Yoan Pechaud, David Huguenot, Giovanni Esposito, Gilles Guibaud, and Eric D. van Hullebusch

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In Situ Chemical Reduction of Chlorinated Organic Compounds . . . 283 Romain Rodrigues, Stéphanie Betelu, Stéfan Colombano, Theodore Tzedakis, Guillaume Masselot, and Ioannis Ignatiadis

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The Nature of Manganese Oxides in Soils and Their Role as Scavengers of Trace Elements: Implication for Soil Remediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 399 Sylvain Grangeon, Philippe Bataillard, and Samuel Coussy ix

List of Contributors

Olivier Atteia Bordeaux INP – ENSEGID, Environnement – Carnot ISIFoR, Pessac, France

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Philippe Bataillard BRGM (French Geological Survey), Orléans, France Stéphanie Betelu BRGM (French Geological Survey), Orléans, France Grégory Cohen Bordeaux INP – ENSEGID, EA 4592 Géoressources Environnement – Carnot ISIFoR, Pessac, France Stéfan Colombano BRGM (French Geological Survey), Orléans, France Samuel Coussy BRGM (French Geological Survey), Orléans, France Hossein Davarzani BRGM (French Geological Survey), Orléans, France Giovanni Esposito Department of Civil, Architectural and Environmental Engineering, University of Napoli “Federico II”, Napoli, Italy Nicolas Fatin-Rouge Institut UTINAM, UMR CNRS 6213, Université de Bourgogne Franche-Comté, Besançon, France Sylvain Grangeon BRGM (French Geological Survey), Orléans, France Dominique Guyonnet BRGM (French Geological Survey), Orléans, France Patrick Höhener Laboratoire Chimie Environnement (UMR 7376), Aix-Marseille University - CNRS, Marseille, France David Huguenot Laboratoire Géomatériaux et Environnement, Université Gustave Eiffel, Champs-sur-Marne, France Ioannis Ignatiadis BRGM (French Geological Survey), Orléans, France Florie Jousse Bordeaux INP – ENSEGID, Environnement – Carnot ISIFoR, Pessac, France

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List of Contributors

Guillaume Masselot ADEME (French Environment and Energy Management Agency), Angers, France Sagyn Omirbekov BRGM (French Geological Survey), Orléans, France Yoan Pechaud Laboratoire Géomatériaux et Environnement, Université Gustave Eiffel, Champs-sur-Marne, France Romain Rodrigues BRGM (French Geological Survey), Orléans, France LGC (Chemical Engineering Laboratory), Toulouse, France ADEME (French Environment and Energy Management Agency), Angers, France Theodore Tzedakis LGC (Chemical Engineering Laboratory), Toulouse, France Eric D. van Hullebusch Institut de physique du globe de Paris, Université de Paris, Paris, France Clément Zornig BRGM (French Geological Survey), Orléans, France

Chapter 1

Contaminant Mobilization from Polluted Soils: Behavior and Reuse of Leaching Solutions Nicolas Fatin-Rouge

Abstract Soil and water contamination by toxic elements or molecules cause risks of chemical pollutions. These pollutions can heavily affect resources and activities of humans and ecosystems. By nature, the remediation of environmental pollutions is a constraint, because its application is hampered by high costs. Soil and groundwater remediation implements a variety of strategies, technologies, and practices to face the diversity and the complexity of every cases. Contaminant recovery is among the three main strategies implemented for the remediation of contaminated soils, and contaminant separation or mobilization by water is widely used for this purpose. As water solubility is the driving force of dissolution, selective chemical agents are used to enhance the mobilization of a wide range of contaminants in aqueous solution from permeable soils. This chapter provides the basic principles about the distribution of contaminants between the soil compartments and the transfer kinetics between these pools. Extracting agents, technologies, and methods used to mobilize priority contaminants are reviewed, considering performances, benefits, costs, and risks within a sustainable approach of reuse and of waste recycling. On the basis of an analysis of the existing literature, the possibility to reuse effective extracting agents and solutions is discussed, considering the overall performances, the costs of the separative methods implemented as well as waste management. However, multicontamination, restrictive regulations or specifications, availability of appropriate installations and the lack of reported full-scale assessments limit the use of this strategy for the remediation of polluted zones. An example of field test using ultrafiltration for the recovery and reuse of surfactant solutions to flush creosotecontaminated soils is presented. Field tests show that the part of costs dedicated to extracting agents is often minor. In the short term, it is hardly compensated for by capital and operating costs of the separative technologies implemented for extraction solution regeneration.

N. Fatin-Rouge (*) Université de Bourgogne Franche-Comté – Institut UTINAM, UMR CNRS 6213, Besançon, France e-mail: [email protected] © Springer Nature Switzerland AG 2020 E. D. van Hullebusch et al. (eds.), Environmental Soil Remediation and Rehabilitation, Applied Environmental Science and Engineering for a Sustainable Future, https://doi.org/10.1007/978-3-030-40348-5_1

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N. Fatin-Rouge

Keywords Contaminants · Soil · Groundwater · Behavior · Remediation · Extraction · Regeneration · Reuse

List of Abbreviations BP BTEX CB CD CMC COD DNAPL EOR FA GW HA HHV HOC IDSA LB LNAPL MF MTE MWCO NAPL NF NOM NVC OM PAH PCB PH PITT PV RO SC SDS SF SFWRS SL SVC SW SZ

Boiling point (
C) Benzene, toluene, ethylbenzene, xylenes Cocamidopropylbetaine Cyclodextrin Critical micellar concentration (mol l1 or %) Chemical oxygen demand Dense nonaqueous phase liquid Enhanced oil recovery Fulvic acids Groundwater Humic acids Highest heating value (MJ ton1) Hydrophobic organic contaminant Iminodisuccinic acid Laurylbetaine Light nonaqueous phase liquid Microfiltration Metal trace element Molecular weight cut-off (Da i.e. g mol1) Nonaqueous phase liquid Nanofiltration Natural organic matter Nonvolatile compound Organic matter Polycyclic aromatic hydrocarbons Polychlorobiphenyl Petroleum hydrocarbons (C10–40) Partitioning interwell tracer test Pore volume of a porous material (m3) Reverse osmosis Solubilization capacity of a micellar surfactant for a given host molecule (mol mol1) Sodium dodecylsulphate Soil-flushing Soil-flushing with reused solutions Soil-leaching Semi-volatile compound Soil-washing Saturated zone

1 Contaminant Mobilization from Polluted Soils: Behavior and Reuse of Leaching. . .

TMP TOC TPH UF USZ VC VOC

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Transmembrane pressure (Pa) Total organic carbon (mg kg1 or %) C10–40 total petroleum hydrocarbon index (mg kg1) Ultrafiltration Unsaturated zone Volatile compound Volatile organic compound

List of Symbols: Capital Letters C CS D Dw H K Kp Koc Kow Km,i Mw Nca P Pc R S Si Ssp Sh0 T V VV W Γi Χi

Concentration (mol l1 or mol m3) Surfactant concentration (mol l1 or mol m3) Dispersion coefficient (m s2) Diffusion coefficient of a pollutant in water (m s2) HENRY constant (Pa m3 mol1) Equilibrium constant or hydraulic conductivity (m s1) Partition coefficient of a solute between two nonmiscible phases Soil organic carbon–water partitioning coefficient Octanol/water partition coefficient Partition coefficient of a contaminant i between the micellar phase and water Molecular weight (Da i.e. g.mol1) Capillary number Pressure (Pa) Capillary pressure (Pa) Retention factor or curvature radius (m) Surface (m2) Solubility of a compound (mol l1) or local saturation of a mobile phase Surface area (m2 m3) Dimensionless Sherwood number Temperature (
C or K) or transmission (%) Volume (m3) Porous volume (m3) Energy (J) Excess concentration of a surfactant i at a o/w interface (mol m2) Mole fraction of a component i

List of Symbols: Small Letters a dp eq g

Activity coefficient (mol l1 or mol m3) Characteristic length of pores (m) Value at the equilibrium Gravity constant (9.8 m s2)

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h k t v xy z β γ ε η μi π ρ θ

N. Fatin-Rouge

Height (m) Permeability (m2 or D) or kinetic rate constant Time (s) Velocity (m s1) or kinetic rate (mol l1 s1) Distance (m) Altitude (m) Stability constant of a complex Interfacial tension (N m1 or mJ m2) Porosity Dynamic viscosity (kg m1 s1) Chemical potential of a component i Osmotic pressure (Pa) Density (kg m3) Contact angle (deg. or rad.)

List of Symbols: Subscript P R T 0 ads d m o s w

Permeate compartment Radius of curvature (m) or feed compartment or retentate Total Initial value Adsorbed phase Dissolved Micellar phase of surfactant Nonaqueous phase liquid nonmiscible with water Solid phase Water phase

1.1

Context

1.1.1

Treatments of Polluted Soils

Considering the protection of people and living organisms against the impacts of pollutions by toxic substances, the management plans of these situations aim at eliminating first the source zones which generate these pollutions and then to prevent further contaminants transfer toward biological targets (Page 1997; Testa and Winegardner 2000). The management methods and treatment strategies of source zones depend on the local context, i.e., the time constraint, the accessibility to the source zone, the watersaturation level of the ground, the type and concentration of contaminants, the planned use after treatment and the concentration-levels or leaching potential to be

1 Contaminant Mobilization from Polluted Soils: Behavior and Reuse of Leaching. . .

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Reused Water Polluted Soil

Active Agents

Coarse Fraction Coagulant / Polymer Sludges

Reused Water Air

Sludges

Fig. 1.1 Pictorial working principle of soil-washing (SW) processes

reached. The management methods cover ex situ, on site, or in situ remediation activities, whereas the strategies on risks elimination in relation to source zones are the contaminant recovery, the immobilization, and the degradation in innocuous metabolites (US Department of Energy 1999). The ex situ management method is traditionally used because it considerably reduces the uncertainties for reaching the remediation objectives. It consists in excavating the polluted zone before transporting it into a secured landfill site or a processing plant. Limitations for storage in landfill are, on the one hand, the regulatory requirements regarding the type of materials, the contaminant concentrations, and their leaching potential and, on the other hand, the Nimby factor. The processing units implement thermal treatments (i.e., vitrification, incineration, desorption) or immobilization (stabilization and solidification) or degradation (biopiles) or washing, which depend on characteristics of the material (e.g., nature, moisture content and type and concentration of contaminants, Khan et al. 2004). Washing units are similar to those used for mining of ores and water is primarily used to mobilize the targeted compounds from fines (lower than 63 μm), since the majority of pollutants are immobilized by adsorption or capillarity (US EPA 2002). They implement crushing (optional), sorting (lower than 100 mm and magnetic sorting for the removal of macroscopic metallic pollutants), water-enhanced soil disintegration, screening between 1 and 100 mm, sieving between 0.04 and 1 mm and cycloning of the polluted sludge, coagulation and flotation to remove dissolved and colloidal fractions of pollutants (Zhang et al. 2001; Dermont et al. 2008). All these steps allow reducing the amount of leached soil and facilitate the accessibility to pollutants and their recovery from soil particles. Chemicals are added to improve the physical processes, especially surfactants. Figure 1.1 shows a scheme about this approach. This treatment method is preferentially suitable for unsaturated

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heterogeneous materials, in which an important fraction of pollutants are macroscopic nonexplosive solids (e.g., bullets in soils from shooting galleries, aggregation of soil particles and tars) or soiled building materials. The on-site management method uses mobile soil washing systems. It decreases the cost and the contaminants dissemination risks associated with the transport (estimated to 10 € ton1 each 100 km) of the polluted materials. It requires low energy technologies easy to implement, providing an effective and reliable treatment. On the one hand, in washing systems of excavated materials, the thorough contact with the washing fluid may cause important losses of active agents by adsorption onto the polluted material. In addition, materials with high content in organic carbon (larger than 10%) and high cation-exchange capacity may cause problems. On the other hand, most of in situ treatments are unable to destructurate ground which makes the contaminant recovery more difficult. Indeed, soil-flushing (SF) treatment is much more challenging, mainly because strongly limited by the accessibility to the polluted zones, then by the monitoring of its efficiency and finally by the contaminant recovery. Nevertheless, they avoid the release of contaminants in the atmosphere and thereby reduce potential contamination risks. Therefore, a geologic and hydrogeologic study of the site, as well as a selection of technologies and operating conditions through treatability studies performed at laboratory scale are essential, but they delay the start of fieldworks. Besides cost and efficiency, the choice for treatment technologies is based on the risks associated to their implementation (Onwubuya et al. 2009; Caliman et al. 2011). Physical treatments that contain or transfer contaminant from a compartment to another are preferred, because of their robustness and the absence of chemical additives being themselves dangerous (e.g., potentially hazardous for handlers, secondary pollutions, uncontrolled mobilization of contaminants). Soil leaching (SL) for pollutant recovery is implemented when physical technologies are not adapted, because of contaminants properties (Sect. 1.2.1) or planned land use. A scheme for SF is shown in Fig. 1.2. Its feasibility is limited by the soil matrix permeability, which should be higher than 105 m s1. Soil fracturation (pneumatic or hydraulic) or jet slurring may be used to improve the flow control in low permeability zones (Thiruvenkatachari et al. 2008). Its efficiency is limited by permeability contrasts within the treated zone and thus it is necessary to ensure that the most permeable and slightly contaminated zones are blocked, in order to concentrate the action on the most contaminated ones (Hirasaki et al. 1997). Hence, geological and hydrogeological site characterizations are required. Such studies focus on water permeability, porosity, particle size distributions, heterogeneity, organic carbon content, and cation-exchange capacity. The extraction kinetics of targeted pollutants plays a major role in SF (Sect. 1.2.2). Because water is often inefficient to mobilize contaminants that are strongly bound to soil, specific additives can be added to improve the treatment effectiveness (Sect. 1.3). However, the dilution of contaminated leachates by groundwater (GW) is also a problem for their treatments, because the latter are generally less efficient and cost-effective on diluted pollutions and the large volumes to deal with. The treatment strategy consists in reducing the mass of pollutant and starts with high productivity technologies (e.g.,

1 Contaminant Mobilization from Polluted Soils: Behavior and Reuse of Leaching. . .

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Fig. 1.2 Schematic illustration of soil flushing

pure phase recovery or excavation) before the treatment of the residual fraction. It is on this level that SF occurs. However, it can be used to enhance pure nonaqueous phase liquid (NAPL) recovery during the last stage of this preliminary treatment. Leaching solutions are injected upstream or within the polluted zone. The injection system involves vertical or horizontal wells, trenches or water-spray equipment. Polluted leachates are recovered using underground drainage systems before being pumped up to the surface level for treatment. Treated wastewater may be discharged in sewer or into the polluted zone. At the end of the treatment, the remaining contaminant in soil and in the treated wastewater should be at the mg kg1 level or less. It is considered that washing may reduce pollutant concentrations up to two orders of magnitude, depending on the selected treatment strategy; for soil-washing (SW) removal yield can reach value higher than 95% and for SF removal yield ranges between 0 and 95% (Atteia et al. 2013). For in situ treatments, the assessment of the pollutant mass still present as residual and its delineation are essential. Sampling strategies that allow for qualitative and quantitative analyses within statistic frameworks are required (Zhang 2007). When the contaminants are hydrocarbons, the partitioning interwell tracer test (PITT) may be advantageously used for that purpose (Burt and Christians 2001; Park et al. 2009). The monitoring of pollutant concentrations in soil leachates and their volumes allows to assess the remaining amount of pollutant on the basis of mass balances analyses. Pollutant concentrations and their leaching potential are both considered since the risks are related to the mobile fraction (ter Laak et al. 2007). SL is attractive, because it enables a treatment of soils displaying pollutant concentrations as high as tens of

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thousands of mg kg1 at relatively low costs, ranging from 20 to 220 € ton1, depending on the technology implemented (Khan et al. 2004; BRGM 2010).

1.1.2

Why the Leaching Solutions Should Be Reused?

Water recirculation is commonly used in SW since every kg of soil treated requires several liters of water (Zhang et al. 2001; Lee et al. 2005; Elgh-Dalgren et al. 2009). The reuse of soil leachates should be considered as it is required to recover and treat highly contaminated wastewater. It is stimulated in the current context of sustainable development and economic optimization. Water reuse is highly encouraged for two reasons: on the one hand, in order to reduce water withdrawals from the natural environment and water stress on aquatic populations, and on the other hand, within a strengthening regulatory context, to reduce the fluxes of contaminant discharged into the natural environment (http://seee-cms-rec.eaufrance.fr/s-informer/). The zeroliquid discharge in receiving natural environment is now promoted (Qurie et al. 2015; Tong and Elimelech 2016). Besides, the reuse of chemical additives present in soil leachates allows the impacts of the raw materials extraction to be reduced, since the transport and transformation steps to supply these chemical agents may be avoided (Sect. 1.4). In fact, there are minimal concentrations of active chemicals, often substantial (e.g., critical micellar concentration (CMC) for surfactants when used for contaminant solubilization) below which additives have no significant role, and, actually, the reuse of these molecules helps to fill this gap. The use of more and more sophisticated molecules subjected to restrictive specifications (e.g., biotoxicity, biodegradability, and specific activity) justifies the economic interest of their reuse. The cost of most of usual active agents is ranging from hundreds to thousands of euros per ton. Mass balances about the transfer of chemicals between the top soil and the underground reveal four main fractions: sorbed, (bio)degraded, lost in groundwater and recovered at the ground surface. Figure 1.3 presents the evolution of these

1

Fraction

Recovered

Biodegraded Sorbed Lost in GW 0

Time (a.u.)

Fig. 1.3 Schematic illustration of the fate of chemicals during soil flushing. GW groundwater

1 Contaminant Mobilization from Polluted Soils: Behavior and Reuse of Leaching. . .

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four fractions along the main phase of SF operations. Therefore, a preliminary study at lab scale is required to select the suitable additives and the conditions of their use, in order to minimize their loss caused by the first two processes over the implementation period (Ahmed Mohamed et al. 2013; Efligenir et al. 2013). Finally, as sustainable development goals aim at supporting a cleaner production approach, pushing away the concept of ultimate waste by supporting the idea to reintroduce these materials into the economy (Dondi et al. 1997). For on-site treatments, the treated soil and the building materials are usually used to fill the excavated zones; however, the latter start to be reintroduced onto the building materials market. About the valuation of contaminants, the market price of the most usual metals and metalloids recovered in brownfields is ranging from 1.3 to 18 € kg1 (CGDD 2013). Electrolysis is often used for their simultaneous separation and purification (Fedje et al. 2013). Recovered hydrocarbons are managed exclusively by thermal mineralization in specific treatment units usually located outside the treated area. Nonhalogenated compounds (halogen level 10 Pa or Henry constants (H ) > 100 Pa m3 mol1 (Zhang 2007). They are mainly removed by desorption technologies (thermal or air-driven, i.e., venting or sparging; US EPA 1997). SVCs evaporate slowly at room temperature, including polycyclic aromatic hydrocarbons (PAHs), phenols, polychlorobiphenyls (PCBs), and dioxins. SL competes with contaminant degradation and immobilization. Both of them are limited by the feasibility (e.g., contaminant recalcitrance), the need for groundwater monitoring, even groundwater recovery and treatment when toxic and persistent metabolites are produced. Recalcitrance of contaminants is defined as their persistence in the environment because of the absence of degradation of the element or the molecule considered. It may be the result of special chemical stability, high toxicity, very low water solubility, or because of local hindrance which prevents the access to the active sites of the molecule (Linde 1994; Boethling et al. 2007). It is considered that a compound is poorly or highly soluble depending on if its concentration in solution is lower than 0.15 or higher than 10 g l1 at room temperature. Chemical pollutants are characterized by their density and volatility at least and they may be divided into five main categories (Zhang 2007; BRGM 2008): 1. Materials (e.g., asbestos, NAPLs): they are physically separated. 2. Metallic trace elements (MTEs, the most frequent are Ag, Bi, Cd, Co, Cr, Cu, Hg, Mn, Ni, Pb, Sn, Tl, V, Zn and radioactive actinides such as Pu, U) which cannot be degraded; they are immobilized or mobilized (Wuana and Okieimen 2011). Elemental MTEs have specific gravities often larger than five, which favor their removal on the basis of sedimentation rates. Their volatility is very low, excepted for the alkyl forms (e.g., BP for tetraethyl lead is 80
C) and for the elemental form of Hg (11.8 Pa at 80
C). Like for Hg, the alkylation may be the result of a natural process with aging. In aqueous solution, their speciation, and therefore their behavior, depends on their oxidation state and on the pH. Often MTEs are in the cationic form, but some elements at oxidation state  IV are under anionic forms (e.g. Bi, Cr(VI)) and some other like Ni may be present as zero-valent colloids (Gillow and Hay 2016). An example of complex speciation is the one of uranium: whereas the uranyl ion (UO22+) is among the major soluble species of this element in oxic acidic to neutral pH conditions in absence of alkalinity, there are numerous species in GW due to a potentially high alkalinity and the presence of O-, N-, P-, or S-donor ligands (e.g., UO2(CO3)22, UO2(CO3)34). Most of the soluble anionic species of elements present in soils are usually treated by immobilization; thus, it is quite inappropriate to consider them within the framework of this theme. Trivalent MTE have often relative low mobility in the environmental

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N. Fatin-Rouge

conditions for both solubility and kinetic reasons. It is noteworthy that MTE immobilization is not always a suitable treatment, as for example when top soils are contaminated by radioelements or when stabilized in the form of strong soluble complexes. Except for sulfide under anaerobic conditions and oxide compounds, most of oxidized MTE are mobile. For example, in France their regulated concentrations in water for discharge into the natural environment is ranging from 0.1 to 5 mg l1 for the non-radioactive elements (Rodier et al. 2005). In conclusion, the most reactive oxidation states follow usually the order M(I) > M (II) > M (IV), and so on, and the species that could be mobilized are essentially mono and divalent MTEs (Wilkins 1991). 3. Metalloids (the most frequent are As, Sb, Se and Te) which have a rather low volatility except for alkylated compounds of As and Se. Metalloids are dissolved in water as oxyanions and are often present in soils as low solubility salts with divalent or trivalent cations. They can be mobilized from soils by competing anions such as phosphate or carbonate. Like for MTEs, their speciation often depends on pH and redox potential. For example, As exists in the trivalent (arsenite H3AsO3) and pentavalent (arsenate H3AsO4) oxidation states in GW and form oxyanion complexes of different charges and properties upon pH; up to height different forms are possible in pure water that behave differently in GW (Watt and Le 2003). In addition, under anaerobic sulfidic conditions, As may be found as soluble thioarsenic complexes. In contrast to arsenite and thioarsenic forms, arsenate is easily removed from GW by precipitation. In France, the regulated levels of metalloids for water discharged into the natural environment are ranging from 0.1 to 5 mg l1 (Rodier et al. 2005). 4. Salts (the most usual are ClOx¼0–4, CN, F, NOx¼2–3, SO42) which are inherently nonvolatile but very soluble in water. The management strategies are either (bio)degradation by redox reactions or SL combined with on-site treatment of leachates using degradation or concentration using reverse osmosis or adsorption (Nozawa-Inoue et al. 2005; Srinivasan and Viraraghavan 2009; Bardiya and Bae 2011). 5. Hydrocarbons: they are distributed in several categories: light, heavy, saturated, unsaturated among which aromatic (BTEX and PAHs), with heteroatoms (especially halogenated, oxygenated, nitrogen, sulfur and phosphorus compounds). They are characterized by their octanol–water partition constant (Kow) at room temperature, in order to categorize them as hydrophilic (log Kow < 2) and hydrophobic (log Kow > 4) (Schwarzenbach et al. 2003). Besides, the superficial soil layers are often loaded with organic matter (OM). The soil organic carbon– water partitioning coefficient (Koc) of a compound is the ratio of its mass that is adsorbed in the soil per unit mass of organic carbon in the soil per its equilibrium concentration in solution. It is used to differentiate compounds that can be easily adsorbed because of their hydrophobic character (log Koc > 4) from hydrophilic ones (log Koc < 2). Most of the organic contaminants, mostly displaying hydrophobic features, have relatively low water solubility. Nevertheless, polar compounds and especially those able to generate H-bonds with water (e.g., alcohols, amines, halogenated compounds) may have aqueous solubility much higher than

1 Contaminant Mobilization from Polluted Soils: Behavior and Reuse of Leaching. . .

13

1 g l1. For water non-miscible hydrocarbons, their behavior in groundwater mostly depends on their density (1 sinking). By taking events of pollution into account, the French commission for sustainable development has distinguished hydrocarbons in four main categories (CGDD 2013): (1) Usual hydrocarbons (gas, fuel oil) display a density below 1 and very low water solubility, whereas gas is volatile and rather biodegradable, fuel oil is semivolatile and less biodegradable; (2) BTEX are hardly oxidizable monocyclic aromatic VCs with density lower than 1 and a rather high solubility in pure water ranging from 0.17 to 1.8 g l1 (log Kow between 2 and 3). (3) Polycyclic aromatic hydrocarbons (PAHs) are compounds of a large family (hundreds of compounds at least) made of fused aromatic rings that have densities higher than 1. They can have heteroatoms (O, N, S) within their rings that improve their solubility in water. Nevertheless, their water solubility is very low for the most usual molecules (log Kow > 4), except for naphthalene (C10H8) that behaves between PAHs and BTEX; it is considered as volatile and its solubility in pure water is 32 mg l1. They are very abundant in tars and are strongly adsorbed to soils. (4) Halogenated compounds (mainly chlorinated and fluorinated ones) which represent a large family with different behaviors: some of them are VCs and display high solubility in water (e.g., up to 4.5 g l1 for trichloroethylene (TCE) at room temperature, RT), while some others like polychlorinated benzene, PCBs and dioxins, are lipophilic, concentrate in sediments and behave more like SVCs. All these compounds are poorly biodegradable because of their high toxicity. However, the reductive dehalogenation is more and more used for in situ remediation (Stroo et al. 2012). Because of their different behavior, these compounds must be considered individually when treatment is decided. The reality of environmental pollutions is often much more complex, because of the mixtures of substances and their interactions with the soil organic matter, but their characteristics are often attenuated (Mulligan and Yong 2004; Fakour and Lin 2014). The effect of local conditions (e.g., pH, redox potential, permeability, ligand concentrations) on contaminants mobility is difficult to predict. Knowing the speciation of MTEs and metalloids is critical in order to understand the mechanisms that control their mobility. The determination of pollutant speciation is needed in order to elaborate an accurate conceptual site model and suitable treatments. Therefore, the measurement of leaching potentials for contaminants and the soil buffer capacity are necessary before considering a treatment in order to, whenever it is required, make it the least impacting and the most economic possible (Gillow and Hay 2016). In conclusion, SL is of special interest to water-soluble, recalcitrant, and poorly volatile contaminants. Therefore, it is often used for salts, usually without the need for additives, unless a redox treatment would be more appropriate. Besides, the most usual contaminants treated by leaching are metals in mono and divalent oxidation states and low volatile and recalcitrant hydrocarbons (PAHs and low solubility chlorinated compounds such as dioxins and PCBs); however, they mostly require the use of additives in order to improve their water solubility (Sect. 1.3).

14

1.2.2

N. Fatin-Rouge

Overview of the Various States of Contaminants, Mechanisms, and Extraction Kinetics

Soil matrix properties which influence contaminant transport and transfer are porosity, permeability, chemical composition (inorganic and organic components), pH, and redox potential (BRGM 2008). Soil matrix permeability, porosity, and the specific surface area of soil particles play an essential role on the occurrence of contaminant–soil interactions. Contaminants are partitioned between different phases (gas, groundwater, and solid phases). Figure 1.4 presents the various contaminants physical forms encountered in soils. During aging, many poorly volatile contaminants tend to become much less accessible because of their hydrophobicity or strong chemical interactions with the soil material or even because of mineral precipitation (Alexander 2000; Frische et al. 2003). Therefore, the older the contamination is, the lower the treatment efficiency is for the removal of the remaining contaminant fraction. Usually, the finest soil particles with the highest surface areas and strong capillary trapping are those in which the highest contaminant concentrations are usually observed. According to Brion and Pelletier (2005), they play a major role during the adsorption phase, but a little one during the sequestration phase.

Fig. 1.4 Different physical forms of organic pollutants in soils [adapted from Volkering et al. (1997)]: solid particulates (a), liquid films onto soil particulates (b), liquid trapped into micropores (c), adsorbed onto soil particulates (d), adsorbed into micropores (e), and entrapped within soil matrix (f)

1 Contaminant Mobilization from Polluted Soils: Behavior and Reuse of Leaching. . .

1.2.2.1

15

Behavior of NAPLs

NAPLs may be composed of organic, organometallic, or inorganic contaminants and elemental Hg. The conceptual model for the migration of NAPLs is shown in Fig. 1.5. In the unsaturated zone (USZ), NAPLs migrate in depth and spread out under the simultaneous action of gravity and capillary forces as long as the infiltrated amount and pressure ensure the physical continuity of this phase (Ahlers and Dunn 2016; BRGM 2008). In this case, contaminants are in equilibrium with soil gas. Light and dense NAPLs (LNAPLs and DNAPLs, respectively) have a similar behavior in the USZ, which is determined by their viscosity, their volatility, and their interactions with the solid material (adsorption, capillary forces). Key parameters which act on the transport of a NAPL through the soil matrix and in contact with groundwater are its viscosity, its density, and its wettability (Williams and Wilder 1971). When a free NAPL is no more continuous, it is trapped in a residual state, in several forms from pools to small accumulations like droplets or clots. This latter form of residual is easier to mobilize by dissolution in water (Maire and Fatinrouge 2017), because of their higher contact with the aqueous phase (rebound effect for plumes). For liquid phases in a porous environment, the local measure of saturation for any phase i, Si, is defined as:

Fig. 1.5 Conceptual scheme for the migration of nonaqueous phase liquids

16

N. Fatin-Rouge

Fig. 1.6 Contact angle and interfacial tensions between a solid and two nonmiscible fluids

γow

Oil

Water

θ

Solid γos

Si ¼

Vi Vv

γws

ð1:2Þ

where Vi and Vv are the volume filled by the phase liquid i and the porous volume, respectively. The interfacial tension γ, is defined as the energy (N m1 or mJ m2) required to make a unit surface (dS) at the interface between two nonmiscible phases, according to: δW ¼ γdS

ð1:3Þ

Lowering the γ-value promotes NAPL emulsification. Wettability is defined as the tendency of a fluid to spread out on a solid surface in the presence of another nonmiscible fluid. This notion is interesting because of its influence on the distribution of the nonmiscible fluids in pores and its effects on their displacement. The contact angle θ is an important measurement directly related to the surface wettability (Fig. 1.6). It is related to interfacial tensions through the Young equation: cos θ ¼

γ os  γ ws γ ow

ð1:4Þ

where the subscript w, o, and s hold for the wetting and non-wetting liquid phases and the solid phase, respectively. Besides, when two nonmiscible liquid phases are in contact, there is a pressure difference at their interface called capillary pressure (Pc, in Pa), which points toward the wetting fluid when the static equilibrium is reached. Pc is calculated according to the Laplace equation: 

 1 1 þ Pc ¼ Po  Pw ¼ γ cos θ R1 R2

ð1:5Þ

where R1 and R2 (m) are the main curvature radii that characterize the interface. By convention these radii are positive if the center of curvature is on the side of the non-wetting fluid. By considering the concept of average curvature, the Young– Laplace equation is as follows:

1 Contaminant Mobilization from Polluted Soils: Behavior and Reuse of Leaching. . .

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Fig. 1.7 Water rise into a water-wet capillary (left) and water down into an oil-wet capillary (right) q ~ 0º

Oil R

Oil

h h

Water

Water

Pc ¼

2γ cos θ R

q ~ 140º R

ð1:6Þ

where R is the average curvature radius at the interface (m). The displacement of a wetting fluid within a thin pipe (Fig. 1.7) is driven by the capillary pressure according to the Jurin equation: Pc ¼ Δρgh

ð1:7Þ

where Δρ (kg m3) is the difference between the mass density of the two fluids, g is the gravity constant (m2 s1), and h is the height (m). LNAPLs spread out on all the fluctuations of groundwater level and can be partially trapped under the static level because of them. DNAPLs migrate downward to the aquitard as a result of their high densities and they accumulate in zones where the entry pressure (Eq. 1.6) prevents or delays their income in pore throats. Their migration is rate-limited by the pressure to which they are subjected. NAPL trapping in the saturated zone (SZ) arises because of the phenomena of capillary instability (snap-off) and of by-passing (Chatzis et al. 1983). The trapping in soil pores much depends on the ratio between capillary and viscous effects, which is called capillary number (Nca): N ca ¼

ηv γ ow

ð1:8Þ

where η (kg m1 s1) and v (m s1) are the dynamic viscosity and the velocity of the incoming wetting phase, respectively. As shown in Fig. 1.8, residual saturations of non-wetting fluids decrease as the capillary number increases beyond 105. In order to mobilize NAPLs, especially those trapped (droplets, clots, and ganglia), the viscosity and the velocity of the incoming mobile phase (water) must be raised, but first of all, the interfacial tension between water and NAPL must be lowered. Lowering interfacial tensions changes the contact angle at the three-phase interfaces and decreases the capillary entry

18

N. Fatin-Rouge

Fig. 1.8 Variations of residual saturations vs. capillary number [Adapted from Lake (1989)]

pressure. It results in high risk to enhance downward migration of DNAPLs (Rathfelder et al. 2003). However, it can bring benefits when a trapped DNAPL contamination is positioned above a compact and isotropic shallow substratum. Thus, SF can also improve the recovery of pure NAPL, especially when using dynamic GW recirculation (Potter and Killingstad 2016). Another way to enhance NAPL recovery is by increasing the pressure gradient, also called viscous force, within the porous network (Maire et al. 2015). The behavior of chemical mixtures usually differs substantially from the one of pure compounds. A well-known example is coal tars, that are mixtures of many molecules with highly different water solubility and affinity for the solid phase. At the tar/water interface, the least soluble compounds limit the availability of the others (Birak and Miller 2009; Ranc et al. 2016).

1.2.2.2

Behavior of Dissolved Contaminants

This behavior concerns all contaminants, but this section focuses more especially on MTEs ions. Several phenomena are involved which delay or speed up the transport of contaminants in GW. These phenomena are adsorption, precipitation (before or after a change of the oxidation state), (bio)transformation, colloidal transport, and volatilization. Chemical speciation is a determining factor in these phenomena. So, the chemical elements can be found in GW as free ions (e.g., Ca2+, Fe2+), as oxyanions when the oxidation state of MTEs is  IV (e.g., CrO4) and as inorganic complexes (e.g., (UO2)2CO3(OH)3) or organic ones (e.g., Fe(CN)63). Most of MTEs are under the cationic form. These cations easily react with the negatively charged soil phases; they can be precipitated by anions in the form of poorly soluble complexes (e.g., sulfide, carbonates, humic acids (HA), aluminosilicates) or remain dissolved as free ion or being stabilized in water in the form of charged complexes (e.g., with fluvic acids (FA) or malate ligands). The formation of strong complexes

1 Contaminant Mobilization from Polluted Soils: Behavior and Reuse of Leaching. . .

19

follows a two-step process: at first, the formation of weak outer-sphere complexes through electrostatic interactions followed by inner-sphere complexes. The latter are stronger but rate-limited by the dissociation of water molecule from the first coordination shell of the metal ion (Wilkins 1991). The main processes for pollutants mobilization from sources zones are the leaching by means of runoff and infiltrations, the dissolution in water, the volatilization and for soils, the erosion, and the wind transport. Because natural concentrations of extracting agents are usually low in runoff water and in GW, chemical additives are often added to enhance the pollutants mobilization.

1.2.2.3

Contaminant Transfer into the Aqueous Phase

Kinetics of contaminant transfer are limiting for the mobilization. The transfer of contaminants from NAPLs toward the aqueous phase is ruled by the mechanism of dissolution. By considering that the equilibrium of dissolution is reached, the concentration of each NAPL component in the aqueous solution, Ci,w,eq, is given by the Raoult law: C i,w,eq ¼ X i ai Si,w

ð1:9Þ

where Χi is the mole fraction of any NAPL component, ai is its activity coefficient in the NAPL (often assumed to be 1), and Si is its solubility in pure water (mol l1). Although NAPLs are often complex mixtures molecules which follow different dissolution kinetics in the water phase, the use of the Raoult law is common in literature, but in that case, during the NAPL dissolution, the mole fraction of each component is changing like its chemical composition. Once NAPL is immobilized, the transport of pollutants into the water phase is described by the adversion–dispersion–diffusion equation (given in one dimension for simplifying—Abriola et al. 1993):   ∂Sw Cw ∂K w C w ∂ ∂C w ∂S ε ¼ þ εSw Dw  ρo ε o ∂t ∂x ∂x ∂t ∂x

ð1:10Þ

where the subscripts w and o indicate the water and the NAPL phases, respectively, ε is the porosity of the soil, C is the aqueous pollutant concentration (mol m3), S is the saturation of the phase, K is the hydraulic conductivity (m s1), D is the dispersion coefficient of the pollutant (m s2), and ρ is the phase density (kg m3). In Eq. (1.10) it is assumed that the water phase density is constant and that there is no adsorption of pollutant at the solid phase. Considering that the contaminants transfer from the NAPL toward the aqueous phase is limited by their dissolution at this interface, it is described by a first-order kinetic (Yeom et al. 1996):

20

N. Fatin-Rouge

dC i,w ¼ ki Ssp ðC i,ow  Ci,w Þ dt

ð1:11Þ

where ki is the rate-constant for dissolution (corresponding to a mass-transfer coefficient through the interface, in m s1), Ssp (m2 m3) is the specific surface area at the NAPL–water interface, and Ci,w and Ci,ow are the concentrations of any component in the aqueous phase and at the NAPL–water interface, respectively. Considering that the dimensionless Sherwood number Sh represents the ratio of the convective mass transfer to the rate of diffusive mass transport, the mass transfer rate coefficient kiSsp may be related to the hydrodynamic conditions through the modified Sherwood number Sh0 , the characteristic length of the pores (dp), and the diffusion coefficient of the pollutant in water (Dw) according to (Schaerlaekens et al. 2000): k i Ssp ¼

Sh0 Dw d2p

ð1:12Þ

The contaminant dissolution is affected by numerous parameters, such as aging, ionic strength, pH, GW flow, and concentrations of leaching agents. The effects of some of them are detailed for hydrophobic organic compounds (HOCs). Through aging, the shape of source zones changes as they rearrange and break (Stroo et al. 2012). The different microenvironments shown in Fig. 1.4 lead to multi-exponential kinetics for the dissolution of single contaminant (Johnson et al. 2001; Jonker et al. 2005). Moreover, because of deposits during soil morphogenesis or very strong interactions with soils, there is a sequestered fraction of contaminants that cannot be recovered unless the soil matrix is dissolved. pH change in soil pores can modify the partitioning of contaminants through dissolution/precipitation equilibria that involve natural organic matter (NOM, e.g., contaminant transport) or contaminants when they are weak electrolytes (e.g., phenolic compounds), or through changes in the electrostatic interactions between pores surface and charged contaminants. The water flow velocity through soil pores changes the dissolved contaminants concentrations in two ways: on the one hand, by modifying the average contact time between water molecules and adsorbed contaminants and, on the other hand, by changing the thickness of the interfacial layer between NAPL and water. As the flow rate increases, the thickness of the interfacial layer decreases and the mass transfer coefficient toward the water phase increases according to Eqs. (1.11) and (1.12). Moreover, as discussed in Sect. 1.2.2.1, it can enhance the mobility of NAPLs as the Nca value increases. Several authors have shown that contact time has a critical effect on dissolved pollutant concentrations; the latter may be increased several times after several hours of contact between the extracting agents and pollutants in soil pores (Pennell et al. 1994; Schaerlaekens et al. 2000; Rathfelder et al. 2001; Taylor et al. 2001).

1 Contaminant Mobilization from Polluted Soils: Behavior and Reuse of Leaching. . .

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Finally, the dissolved organic matter (OM) may significantly increase contaminants solubility, especially for HOCs, MTEs, and metalloids. This OM originates from natural processes of decomposition (FA and HA) and/or anthropogenic activities. For HOCs, adsorption onto mineral soil is low except for clay or when organic carbon represents a significant fraction (Karickhoff et al. 1979; Ranc et al. 2016). The partition of a contaminant between pore water and OM often obey to a simple partition equation: C o,eq ¼ K p C w,eq

ð1:13Þ

where Co,eq and Cw,eq are the equilibrium concentrations (mol l1) of a contaminant in the organic phase and in the water phase, respectively, and Kp is the partition coefficient of the contaminant between the two phases. OM may be mobilized under some conditions and the apparent contaminant solubility becomes:   Si,app,eq ¼ Si,w,eq 1 þ K p,i Cod

ð1:14Þ

where Si,w,eq is the contaminant solubility in pure water (mol l1) and Cod is the dissolved OM concentration (mol l1). Equation 1.14 can also be applied when a water soluble ligand is added in order to increase the contaminant solubility. In that case, Kp is the apparent equilibrium formation constant for their soluble complex and Cod is the free ligand concentration. Nevertheless, the dissolution ability of OM varies a lot depending on the nature of the soluble complex.

1.3 1.3.1

Contaminant Leaching in Water General Considerations

SL should be considered only when the natural attenuation is not effective enough with respect to timeframe delays for site reuse and when the use of water alone has only a low impact on the amount of mobilized pollutants. In addition, when remediation measures are required, SL performances should be compared to degradation and immobilization ones on the basis of both the conceptual site model, characterization and treatability tests. When SL is selected, it should be carried out with the lowest amounts of additives. The natural contaminants mobilization by GW is usually very slow and the removal of source zones is estimated within decades to centuries (ITRC 2002). In contrast, when the contaminants inertness prevents their reaction with soils (e.g., Ni2+), they are rapidly pulled into GW and treatments address the latter (Evanko and Dzombak 1997). The chemical weathering of rocks and soils is a natural process assisted by biota that involves acids and organic molecules. In practice, the additives used to favor the leaching of contaminants depend mainly on their physicochemical properties. Ex

22

N. Fatin-Rouge

situ and in situ treatments obey different rules. However, the purpose of the treatment should be to remove the mobilizable fraction of contaminant from the contaminated material. Generally, the total contaminant concentration after washing is not predictive of the leaching behavior of the residual contaminant (Tsang et al. 2013). The aim of leaching treatment is to reduce contaminant concentrations below regulatory levels if any, but also to meet the criteria for contaminants leachability of the treated materials considering the different exposure scenarios (Engelsen et al. 2010; van der Sloot and Kosson 2012). For SW, gravity and magnetic separations are operated in order to eliminate macroscopic contaminants, whereas chemicals are mainly involved for fine particles treatment (Dermont et al. 2008; Mann 1999). The soil colloidal fraction, which is the most polluted one, is separated from the solid phase using cheap molecules that adsorb strongly onto surfaces, like citrate or surfactants. In contrast, in SF, the nonspecific adsorption of reagents is avoided and they are used at low concentrations in order to prevent the decrease of the soil hydraulic conductivity by pore clogging and the risk for secondary pollution (Kedziorek and Bourg 2000; Hauser et al. 2005). The use of extracting agents at low concentration requires selectivity and efficiency, which are obtained on the basis of specific effects. They are due to a high molecular organization which favors the formation of contaminant-rich mobile particles or water-soluble complexes.

1.3.2

Organic Contaminants Extraction

In spite of the high efficiency of organic solvents for the extraction of HOCs, they were gradually banished from most of remediation processes for the benefit of the water-based solutions, because of potential health risks and impacts on the environment (Birak et al. 2011; Gomes et al. 2013). Despite vegetal oils being used as organic solvent substitutes, it is rather in the form of oil in water emulsions (von Lau et al. 2014). Besides, supercritical carbon dioxide is used in dedicated extraction units (Fu and Matthews 1999). The following discussion focuses on water-based leaching approaches. Many natural molecules have been assessed, starting with HA and FA, more especially dedicated to HOCs (Molson et al. 2002; van Stempvoort et al. 2002) and their chemically modified parent compounds with enhanced tensioactive properties and stability in aqueous solution (Garcia-Diaz et al. 2015). Cyclodextrins (CDs) are pre-organized natural molecules which offer a hydrophobic cavity. β-CD and its alkylated parent compounds, such as methyl-β-CD whose performances have been improved through hemi-synthesis, show high capacities to dissolve 3–4 rings PAHs in water (Rekharsky and Inoue 1998; Badr et al. 2004; Viglianti et al. 2006). These molecules have low affinity for soil material and usually they are stable for few weeks in saturated soils. Nevertheless, they are too selective and too expensive to be used for soil remediation. The best suitability is obtained for surfactants, which are much used for soil remediation (Paria 2008). They may be of natural origin like for example Saponin. They can act either in the form of monomers to mobilize NAPLs, micelles to dissolve them in water or emulsions when they are

1 Contaminant Mobilization from Polluted Soils: Behavior and Reuse of Leaching. . .

a

23

b Air Water

SHOC

Air Water

SHOC,W

CMC

CMC Log CS

Cs

Fig. 1.9 Variations for the air–water surface tension (a) and aqueous solubility of hydrophobic organic contaminants (b) as a function of the surfactant concentration

used to stabilize the dispersion of vegetal oil droplets for example (Rosen and Kunjappu 2012). Surfactants are amphiphilic molecules displaying at least one hydrophilic head and one hydrophobic tail. Their structural characteristics vary a lot, but they are categorized as charged (cationic or anionic) and neutral (nonionic or zwiterrionic). They adsorb to interfaces and orient in order to lower the o/w interfacial energy γ ow (Eq. 1.3) and they enhance NAPL mobility (Eq. 1.8). The Gibbs equation links its variation to changes in chemical potential of any component of the system, μi (J mol1): dγ ow ¼ 

X Γi dμi

ð1:15Þ

i

where Γi (mol m2) is the surface excess concentration of the ith component of the system. When equilibrium is reached: dμi ¼ RTd ln ai

ð1:16Þ

where ai is the activity of any component of the system. For the simplest case of diluted solutions of a neutral surfactant (i.e., CS  102 mol l1), Eq. (1.15) becomes: dγ ow ¼ RTΓS d ln C S ¼ 2:303RTΓS d log CS

ð1:17Þ

where CS is the surfactant concentration in solution (mol l1). For the air/water interface, the surface excess concentration is constant over a large range of concentration as shown in Fig. 1.9a. Above a fixed value of free surfactant concentration called the CMC, surfactant monomers become no more soluble and associate as aggregates called micelles. In absence of solid phase or NAPL, above the CMC value, interfacial tensions become stable because the surfactant monomer

24

N. Fatin-Rouge

concentration in bulk aqueous solution remains constant. In presence of an adsorbing phase, it happens at a higher value. Through the monomer adsorption at the NAPL– water interface, their interfacial tension is decreased which favors NAPL mobilization. Moreover, some other solution properties change, like the solubility of HOCs which may be significantly increased for CS > CMC (Fig. 1.9b). Indeed, the apparent solubility at the chemical equilibrium Si,app,eq for any sparingly soluble component as a function of CS is given by: Si,app,eq ¼ Si,w,eq þ SCðC S  CMCÞ

ð1:18Þ

where Si,w,eq is the apparent solubility in aqueous solution of the given compound (mol l1) in absence of surfactant, SC is the solubilization capacity (mole mole1), which is specific of the system (surfactant, HOC) and is often constant over a large range of CS (Rosen and Kunjappu 2012). SC is defined as the average number of host molecules solubilized in micelles per surfactant monomer that are under the micellar form. Typical solubilization capacities for PAHs are lower than 0.4, showing that swollen micelles contain more surfactants than HOCs (Zhu and Feng 2003; Liang et al. 2014). Expressed as the partition coefficient, Km,i, for any component i between the micellar and the aqueous phases, it gives: K m,i ¼ X i,m ¼

with

and

X i,m X i,w

ð1:19Þ

C i,m SC ¼ C i,m þ CS,m 1 þ SC

X i,w ¼

Si,w,eq ¼ Si,w,eq V w Cw

where Xi,m and Xi,w are the mole fraction for any solute i in the micellar and in the aqueous phases, respectively, Ci,m and CS,m are the concentrations (mol l1) of the solute i and of the surfactant in the micellar phase, respectively, Cw and V w are the concentration and the molar volume of water, respectively (55.5 mol l1 and 1.8  102 l mol1 at 20
C). Valsaraj and Thibodeaux (1989) have shown that Km,i is related to the partition constant for the solute i between octanol and water, Kow,i, according to the following relationship: log K m,i ¼ α log K ow,i þ β

ð1:20Þ

where α and β are positive values, which depend only on the solubilizing medium. The SC and solubility values of HOCs in micellar solutions decrease as their Kow value. However, low mobilization efficiencies for heavy PAHs larger than four rings are generally observed, because of their very low solubility in pure water, the difficulty to solubilize large molecules in micelles and the longer desorption kinetics.

1 Contaminant Mobilization from Polluted Soils: Behavior and Reuse of Leaching. . .

25

Fig. 1.10 Scheme of the different loci for hosts in surfactant micelles

Therefore, surfactants have an effective multi-action on the mobilization of hydrophobic contaminants. On the one hand, they lower interfacial tensions that contribute to overcome capillary trapping of contaminants in pure phases and that favor HOCs emulsification and mobilization. On the other hand, for CS  CMC in soil pores, they improve the aqueous solubility of contaminant by entrapping them in the form of soluble complexes. The average position occupied by the host molecules in micelles depends on their interactions with surfactants and water. On the basis of experimental measurements, five main positions described in the literature are shown in Fig. 1.10. The most saturated HOCs are in position 5. In contrast, aromatic rings such as benzene stand in positions 2–5 depending on the surfactant structure (Nagarajan et al. 1984), because of their polarity due to the resonance of π-electrons. Micelles swell until they become oil in water emulsions as host molecules incorporate into their hydrophobic lobes. Emulsions are categorized as microemulsions (size Zn (Irving and Williams 1953). Nevertheless, there might be an entropic contribution (T  ΔS
), sometimes major, especially for the formation of complexes with chelating agents. For organometallic complexes, whenever it is possible to make rings of bonds, the number of bonds involved in the ring has a major influence on the complex stability. Five-bond ring complexes are the most stable, whereas below this value the ring stress is too high and above eight bonds there is no stability enhancement. Chelating agents are designed on this observation. Ethylenediaminetetracarboxylic acid (EDTA)-derived polyaminocarboxylates combine the chelate effects of ethylenediamine and aminoacetate moieties (Knepper 2003), offering a number of electron-donor atoms close to the coordination number

1 Contaminant Mobilization from Polluted Soils: Behavior and Reuse of Leaching. . .

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of the metal ion (Lestan et al. 2008). Thus, it is possible to get very stable complexes whose ligand flexibility enables both for high formation and dissociation kinetics that are required for fast and effective metal extraction and then ligand regeneration. Thus, chelating agents stabilize easily MTEs in solution in the form of ions, pulling precipitation equilibria toward dissolution. EDTA has been largely assessed as it offers among the best cost-to-performance ratios (Nowack et al. 2006). Nevertheless, it is so powerful that it frees a portion of the strongly bounded fraction of contaminants, then redistributed between the mobilized and the available fractions (Udovic and Lestan 2009; Zhang et al. 2010a, b). Through the number of e-donor atoms in their backbone, the diversity of EDTA-derived chelating agents offers the ability to modulate both the binding constant of the metal complexes and their dissociation rate, their selectivity toward MTEs and to obtain a strong activity at low concentration. These ligands may also release anionic contaminants, like high oxidation state MTEs and metalloids, through the dissolution of their counter ions showing a high activity potential (Mandiwana 2008). Many recent studies have focused on the biodegradability and the selectivity enhancements of chelating agents in order to overcome the deficiencies of EDTA and parent compounds (Svenson et al. 1989; Nowack et al. 2001; Nowack 2002). The most suitable chelating agents currently available seem to be the iminodisuccinic acid (IDSA) and the ethylene diamine N, N0 -disuccinic acid (EDDS) (Tandy et al. 2004, 2006; Hauser et al. 2005; Ahmed Mohamed et al. 2013; Ferraro et al. 2016). As a general rule, adding aqueous NOM to contaminated matrices generates an increase or a decrease of contaminant leaching depending on its solubility (Molson et al. 2002; van Stempvoort et al. 2002; Moreno-Jimenez et al. 2013). The low solubility of HA is strongly reduced in presence of MTE cations. Besides, an important mobilization of As in the liquid and gaseous phases was generally observed (Beesley et al. 2014). It was also observed that for polycontaminated soils, the HOC removal was enhanced by the synergistic mobilization of metal by chelating agents such as EDTA (Subramaniam et al. 2004; Ehsan et al. 2007; RiveroHuguet and Marschall 2011). Such phenomena can be explained by the high affinity of these organic contaminants for NOM (Ranc et al. 2016). In contrast with strong chelating agents such as EDTA or EDDS, the use of humic substances enabled a reduction of leached concentrations without destabilizing the strongly bounded contaminants. However, the reduction in some MTEs in the solid matrix or in As leachability may be insufficient to meet the regulatory criteria (Tsang et al. 2013). The dissolution mechanism of inorganic and organic contaminants from solid materials involves several successive elementary steps: (1) the transfer of contaminants between binding sites, (2) the chemical reactions, (3) the desorption of contaminants, and (4) the mass transport into the bulk solution. When dissolution rates for soils are slow, the rate controlling mechanisms are either the mass transfer of contaminants within the solid phase or their surface process-controlled detachment. In these cases, the dissolution kinetics follow a zero-order rate law when the steady state conditions at the surface happen:

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N. Fatin-Rouge

v¼

dC ¼kS dt

ð1:24Þ

where k is the rate constant (mol l1 m2 s1) and S is the surface area of the solid (m2). One the one hand, outer sphere surface complexes show small effects on dissolution rates, while the oxidation state of the metal ion remains unchanged. On the other hand, the surface reactivity of inner sphere complexes with ligands is enhanced or inhibited depending on the nature of the ligand. When surface metals cations are bound by water, hydroxide or ligands bound to one metal center, the dissolution is easy. Alternatively, when ligands are bridging several metal centers (e.g., phosphate) or when they are blocking the surface by hydrophobic moieties (e.g., HA, polymers), the dissolution is inhibited. The electron density or negative charge of ligands that enter into the metal coordination sphere decreases its Lewis acidity and enhance the lability of the metal-surface bonds, favoring metal dissolution. The formation of surface complexes is fast, while the release of metal ions in solution is a slow step. There are two simultaneous ways to enhance the solubility of metals ions: (1) the free metal ions concentration is lowered in solution since metals are bound to ligands as soluble complexes and slow surface dissolution and interlattice reactions happen to maintain equilibrium concentrations and (2) the adsorption of a ligand onto the solid surface leads to the formation of ternary complexes as an activated intermediate before the release of a soluble complex and a free site occurs at the solid surface. According to the activated complex theory and assuming the steady-state, the dissolution rate v increases with the surface concentration of the precursor of the activated complex: v¼

dC ¼ k0  C dt

ð1:25Þ

where k0 is the rate constant (m2 l1 s1) and C is the surface concentration of the precursor (mol m2). Thus, there is an ambivalent role of adsorption for ligands as it is involved in surface dissolution, but their use has to be kept low and specific because of costs and preservation of soils properties. Therefore, the balance between surface complexation and metal dissolution (e.g., citrate vs. EDTA) is critical when considering ligands. Thus, the appropriate selection of ligands and process operating conditions are essential both in order to ensure the success of the treatment and to control side effects. This choice builds on a large literature complicated by the wide variability of environmental conditions. It must be supported by lab tests for matrix characterization (especially leaching tests) and treatability. Selectivity of ligands for contaminant removal is important. It may have a kinetic or a thermodynamic origin. The kinetic origin is related to contaminant reactivity and hence to site reactivity also if any. When exchange kinetics are fast, selectivity is imposed by the relative stability of contaminant–ligand complexes with respect to those with competitors and relative concentrations. The selection of the active agents and their concentrations is built on

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toxicity for the receiving medium, biodegradability and efficiency under the planned conditions of application. The measurement of the efficiency includes the minimal loss of ligand in the treated medium (e.g., due to adsorption and ion-exchange interactions), the lowest concentration for which activity is observed, the selectivity for leaching, and the effectiveness over the whole targeted population. As a general rule, the higher the loss is, the higher the matters are (e.g., cost for reagents, pore clogging, uncontrolled release with rebound effect). Nevertheless, the enhanced solubilization of contaminants at field scale is ligand intensive. A typical example of this is the HOCs mobilization with surfactants. Typical SC values  0.27 for the dissolution of HOCs in micelles show that these molecular structures contain much more surfactant than contaminant (Zhu and Feng 2003). Despite this ratio being inverted for emulsions, swollen micelles are favored for many contaminant residuals because of both the limiting HOCs desorption kinetics and the high value for the surfactant to HOC ratios usually used in SL because of the mass transfer law. For this reason, it is desirable that the extracting agents in soil leachates should be recovered for reuse.

1.4

State of Knowledge Regarding Soil Leachate Treatment Technologies that Enable Their Reuse

In order to reuse washing solutions, contaminants present in soil leachates should be removed selectively relative to chemical extractants. In the frame of environmental pollution control, the removal of contaminants prevails over the recovery of chemical agents. The treatment of leachates usually involves a separative process in order to remove and concentrate the mobilized contaminants. Usually, in many processes (e.g., demetallation) ligands are just degraded (e.g., oxidation) to release contaminants which separate from the aqueous phase (Tucker et al. 1999; Finzgar and Lestan 2008; Huang et al. 2016). This chemical-intensive strategy is not sustainable and may lead to secondary pollutions. At the exit of the treated zone, the soil leachate is composed of free contaminant (C), free ligand (L), and their complexes written in the simplified form CiLj. In order to recover and reuse the ligands while separating contaminants from the aqueous phase, two strategies may be implemented: • At least, the selective extraction of C and CiLj while keeping L in treated leachates • Or the selective extraction of C after the fast dissociation of the CiLj complexes Separative treatments processes are classified into physical, physicochemical, and chemical processes. Physical treatments are of special interest because of the absence of added chemicals to the handled solution. The main drawbacks of chemical treatments are the use of chemicals itself, the poor selectivity toward the targeted contaminants, and the potential risks of chemical interferences within the whole process.

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1.4.1

Physical Treatments

1.4.1.1

Air Stripping

It is usually implemented for the removal of volatile contaminants through a straightforward transfer from the aqueous phase to the gas phase. Air stripping is an excellent alternative to distillation, which is not economically feasible. It can be performed in the simplest cases by spraying, bubbling, or mechanical dispersion (e.g., cascade). All the stripping systems work in counterflow to the wastewater in order to increase the amount of contaminant extracted according to mass transfer laws. The use of high surfaces for transfer processes is required. In basic systems, the wastewater flow to be treated is either sprayed at the top of a long column or scattered and uniformly distributed within a packed bed of solid material. However, volatile contaminants are usually organic and surfactants are generally used to extract contaminants from soils. The higher the surfactant concentration is, the lower the removal efficiency is in the vapor phase, because of the enhanced contaminant stabilization in the aqueous phase and then, the increase of the apparent Henry constants (Lipe et al. 1996). Foam production resulting from the decreased surface tension between the air and the water phases by surfactants (Sect. 1.3.2) may hinder a fast fluids circulation in the column. In that case, vacuum stripping of VOCs in nonionic surfactant solutions in co-current mode, or even better membrane pervaporation have been used for the removal of BTEX, naphthalene, and chlorinated compounds (Jiang et al. 1997; Abou-Nemeh et al. 1999; Kim et al. 2007; Topf et al. 2013). Field-tests assessments have been mainly carried out on TCE-contaminated groundwater at low surfactant concentrations and needed ultrafiltration for surfactant recovery before reuse (Sabatini et al. 1998; Vane et al. 2001). The principle of membrane pervaporation is explained in the next section.

1.4.1.2

Membrane Filtration

The filtration of colloids and solutes requires membrane technologies since classical filtration cannot retain particles smaller than few microns. A membrane is a thin selective barrier (about 200 μm thick) separating two compartments which enables the transfer of molecules or colloids thanks to a driving force. The driving force is a gradient between the two compartments either pressure or compound concentration or electric field. The efficiency of a membrane system is characterized by its volumetric flux (m3 s1 m2), its recovery (defined as the ratio of permeate flow to feed flow) and the rejection, R, which is the fraction of molecules or solids that is retained by the membrane and calculated as:

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Fig. 1.12 Scheme of the removal of VOCs from a surfactant solution by pervaporation. P pollutants, V water molecules

  CP R ¼ 100  1  CR

ð1:26Þ

where CP and CR are total concentrations for a molecule or species in the permeate and in feed solutions, respectively. Conversely, transmission, T, is defined as the percentage of solute or solid that is not retained by the membrane. T ¼ 100  R

ð1:27Þ

The membrane configuration refers to the membrane geometry and its position in space in respect to the flows of the feed and permeate fluids. There are four main categories of membrane configurations used: plate and frame, spiral-wound, tubular, and hollow fiber. The choice of a specific membrane configuration is made on the basis of compactness, easiness for cleaning operations that are frequent, and the required cross-flow velocity for the feed flow in order to limit membrane clogging and especially concentration polarization (IAEA 2004). The interest of membrane technologies for the treatments of soil leachates has been recognized for a long time (Sikdar et al. 1998; Das et al. 1999). The selection of the membrane properties depends both on the size of the solutes or particles to be separated, and on the respective affinity of the solutes for the membrane material and the solvent. The smaller the solutes are, the smaller is the pore size of the membrane and the higher the process energy is required. Clarification and desalination refer to the removal of solids and solutes, respectively. The principle of pervaporation consists in transferring molecules in the vapor state from a solution through a dense membrane (Fig. 1.12). The vapor collected in the downstream compartment has a chemical composition different from the upstream compartment because of the membrane selectivity. The mass transfer is

34

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maintained using a lower pressure in the downstream compartment and condensing the molecules onto a cold surface. The transfer mechanism is called “solutiondiffusion” as the first step is the selective solubilization of compounds at the feed compartment/membrane interface that obey to a partition coefficient and the second step is the diffusion-controlled transport within the membrane up to the downstream compartment. The proper choice of the membrane material is essential since it must have a high affinity for the targeted molecules and possibly their mobility within the material must be higher than those of the others to achieve a good selectivity for their transfer. Thus, materials like polypropylene are suitable for hydrophobic contaminants in aqueous leachates. Concentration factors of 1000 are not uncommon. Because of the low solubility of hydrophobic contaminants, an NAPL in contact to a contaminant-saturated water phase is often observed when the permeate vapors are condensed (Jyoti et al. 2015). Reverse osmosis (RO) is so called as it reverses the osmosis, that is the spontaneous diffusion of solvent that takes place when two solutions of different solute concentrations are separated by a semipermeable membrane. Like pervaporation, RO uses dense membranes and follows the same transfer mechanism. The osmosis may be reversed when the pressure applied to the more concentrated solution is higher than the osmotic pressure. The osmotic pressure, π (Pa), is usually estimated assuming diluted solution conditions using the ideal gas law: π ¼ RT

X

ΔC i

ð1:28Þ

i

where ΔCi refers to the concentrations difference between the feed and the permeate compartments for each solute i, R (8.31 J K1 mol1) and T(K) are the gas constant and the temperature, respectively. Osmotic pressures larger than 1 MPa are not uncommon in RO. The pressure imposed upon the feed solution in RO has two components: the one required to overcome the osmotic pressure and the other to flow water through the membrane. As for all the pressure-driven membrane processes, there is a linear relationship between the driving pressure and the volumetric flow to permeate as long as the membrane structure and characteristics remains unchanged. Because of this, in absence of concentration polarization, it is observed that for all the driven membrane processes, the rejection increases with the driving pressure. Therefore, for desalination technologies (RO, nanofiltration (NF)) driving pressures are commonly larger than 1 MPa at least. RO has limited interests in additive/ contaminant separation because the rejection rates are usually larger than 95%. Nevertheless, it is a useful technology for concentration, implemented for zero liquid discharge, since high-quality water recovered in permeates can be reused in many processes (Tong and Elimelech 2016). Filtration Using Porous Membranes Nanofiltration(NF), ultrafiltration (UF) and microfiltration (MF) implement porous membranes of increasing pore size in order to make solute/solute, solute/colloid and solute, or colloid/particulate separation.

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NF membranes are called microporous, because their size (ranging from 0.5 to 2 nm) is similar to that of few water molecules. This process is based on a selective separation between solutes on the basis of steric, electrostatic, and dielectric interactions between the solute and the charged (usually negatively) membrane. The reason for the dielectric effect is the low relative permittivity (about 30) in pores, which is due to exclusion effects on water molecules. Moreover, because of the high surface area of NF membranes, adsorption of solutes may play a major role on their removal at low concentrations. NF is used to separate weakly charged or uncharged molecules from multivalent ions as long as the Donnan effect remains high (e.g., a high charge density onto the membrane and a Debye length larger than the pore size, i.e., ionic strength of solution not high). For example, in NF, transmission ranges usually from 30 to 85% and 1 to 30% for monovalent and divalent ions, respectively (Degrémont 2005). In remediation, NF is often used for the treatment of landfill leachates as it removes efficiently both dissolved total organic carbon (TOC) and MTEs (Chaudhari and Murthy 2010). For the removal of metal ions, NF is often preferred to RO because it has higher permeate flow rates and lower driving pressures (Liu et al. 2008). It has been used for the recovery of oxidants from contaminated groundwater during the treatment of contaminated soils (Liang et al. 2007; Ahmed Mohamed 2014). NF has not been used for chelating agent recovery, since the electric charge carried by free ligands is usually higher than those of their metal complexes and ligands are retained instead of contaminants (Suarez et al. 2013). MF membranes are called macroporous, since their pores, ranging from 0.1 to 10 μm are visible with an optic microscope and are the largest membrane materials. It is used to remove suspended particles using a sieving effect at very low driving pressures (0.03 MPa). In SW operations, it is mainly used to remove solid particles to prevent clogging of NF, RO, or pervaporation membranes (Kujawski et al. 2009). MF is often used to retain degrading bacteria in membrane bioreactors; however, it is very difficult to combine such a treatment with ligand recovery (Degrémont 2005). It has been assessed in order to remove contaminants from oil field-produced water; however, in contrast to UF, it was not able to meet standards for rejection into the environment (Bilstad and Espedal 1996). The main issue with MF in the separation of emulsions from wastewater is that a substantial fraction of contaminants is usually present in microemulsions and in micelles with sizes lower than those of the membrane pores (Peng and Tremblay 2008). UF membranes are called mesoporous, since their pore size ranges between those of NF and MF. It removes colloids and polymers from water and it is used in many industrial applications (Jönsson and Tragardh 1990). It is mainly used in the disinfection of drinking water production or as antifouling treatment before RO, since UF membranes are very effective to remove the suspended matter. The main mechanisms involved in the separation are sieving and adsorption, since a large surface area is contacting the feed solution. These phenomena are responsible for the reduction in permeate flow rates. Highly hydrophilic membranes should be used to prevent the adsorption of hydrophobic molecules. UF is the most interesting membrane technology for oily wastewater treatment, thanks to its high efficiency in oil

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removal and compactness, the absence of chemical additives and relatively low energy costs (He and Jiang 2008). In fact, despite low driving pressures, the major part of the energy consumed is used to maintain a high cross-flow velocity in order to prevent clogging and concentration polarization (IAEA 2004). In turn, this high velocity allows a very simple design for the treatment train (Sect. 1.5). Typically, less than 10% of the filtered volumes are needed for membrane cleaning operations (Degrémont 2005). Irreversible membrane clogging is reduced by using more or less aggressive chemicals such as acids, bases, oxidizing agents, which should be cautiously used (Rabuni et al. 2015). For the hydrophobic solutes in micelles or microemulsions separation, the rejection rates for contaminants in the range 85–99% are maintained all over the range of surfactant concentrations, thanks to adsorption and sieving effect at low and higher concentrations, respectively (Chakrabarty et al. 2010). In contrast, recovery rates for surfactants in filtrates is very dependent on their concentration, solute–membrane and solute–solute interactions (Hanafiah et al. 2018), ranging from about 100% at CS lower than the CMC for hydrophilic membranes to few percents for highly concentrated wastewater (Azoug et al. 1998; Jönsson et al. 2006). UF of contaminated GW during surfactant soil-flushing operations carried out using surfactant concentrations lower than the CMC was investigated for PCBs (Ang and Abdul 1994) and BTEX removal (Sabatini et al. 1998). Surfactant-recovery yields in the range 46–80% were reported. In the latter case, pervaporation was used first in order to remove the volatile contaminants from the contaminated GW, while UF was dedicated to surfactant reconcentration before reinjection. UF has also been successfully tested for the recovery of cyclodextrins from PAHs-contaminated soil leachates using colza oil microemulsions for the PAH retention (Petitgirard et al. 2009). Despite ultrafiltration being widely investigated for the removal of ions and chelated metal using micellar enhanced ultrafiltration (MEUF) for example (Baek et al. 2003; Jung et al. 2008; Rivas et al. 2011; El Zeftawy and Mulligan 2011), few studies have been reported for the treatment of metal complexes-contaminated wastewater with ligand recovery (Zamariotto et al. 2010).

1.4.2

Physicochemical Treatments

1.4.2.1

Sorption

Sorption is a physicochemical process in which a solute becomes attached to a solid. Many materials have been proposed as sorbents, mostly because of their high specific surface area. Well-known sorbents are activated carbon, resins, iron oxyhydroxides, zeolites, but there are many others like biopolymers (e.g., cotton) and waste materials since costs for production and regeneration is a critical parameter. There is a variety of processes through which sorption occurs, e.g., ion exchange, adsorption, absorption, and complexation. The regeneration pathway of sorbents depends on the physicochemical properties of the bound contaminant.

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Sorbents loaded with organic contaminants are regenerated by heating, lowering the pressure or washing with solvents (Sufnarski 1999). When thermal reactivation of the sorbent occurs at temperatures higher than 800
C, a marginal loss of sorbent material may occur for organics. Inorganic contaminants are usually removed in presence of a concentrated solution of competing ions. The selectivity of phenanthrene removal from micellar solutions using activated carbon has been studied in details for several neutral surfactants and their mixtures with sodium dodecylsulfate (Ahn et al. 2007, 2008a, b, 2010). The reported selectivity for phenanthrene adsorption with respect to surfactant was up to 95 and it was enhanced as the fraction of anionic surfactant increased in the mixture of tensionactive agents. In spite of high surfactant recovery yields ranging from 88 to 96% were obtained, often relatively low removal yields for phenanthrene, ranging from 24 to 70%, were reported, but it may reach up to 90% for pure sodium dodecylsulfate (SDS) surfactant. The effect of mesh size for activated carbon particles was studied and it was observed that the selectivity increased with the size of adsorbent. However, the contaminant concentration plays a critical role with respect to removal yield and selectivity and only phenanthrene concentrations as low as 10 mg l1 gave very good results, while PAHs concentrations in soil leachates are rather about several hundreds of mg l1 during washing or flushing operations. Moreover, kinetics for phenanthrene adsorption from micellar wastewater is long, with halflive for fastest conditions of approximately 16 h, since surfactant adsorption is faster. The long time required to achieve equilibrium concentrations and satisfactory removal rates for contaminants is not very practical with respect to residence time in wastewater treatment facilities and surfactant biodegradability. Finally, the cost for activated carbon is similar to the one for surfactant. Besides, Wan (2011) report the selective removal of hexachlorobenzene (80–99%) from soil leachates using activated carbon at 10 g l1 and the simultaneous recovery of a rhamnolipid surfactant ranging from 80 to 90%. Ion exchange using the chelating resins Amberlite IRC 748 has been studied to recover several chelating agents, like EDTA, bound to divalent metals in contaminated soil leachates (Ahmed Mohamed et al. 2013). The treated solutions were reused up to seven times with ligand recovery ranging from 30 to 100%, allowing metal extraction from soils as high as 80–97% using millimolar solutions of chelating agents. It was shown that the low toxicity and highly biodegradable iminodissuccinic acid (IDSA) was suitable for field works since it was completely and shortly recovered, and was very effective to extract the divalent metal ions from the spiked soil. Alcaline-earth cations did not perturb the process because, on the one hand, the selectivity of IDSA and of the chelating resin for MTEs and on the other hand, they were rapidly displaced from chelates by MTEs. The main problem is associated with inert contaminants such as Ni2+, since their removal from soil and then from the extracting agents needs longer contact times which limit the process efficiency. This technology can be easily implemented for wastewater treatment because it does not need any pH modification and it has a low energy consumption. Metal recovery from resins was carried out at pH 2 in the form of concentrated solutions before the electrolytic reduction of free metal cations.

38

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However, the high affinity of the resin for Fe(III) may prevent the use of this technology in iron-rich soils.

1.4.2.2

Precipitation

Neutralization is the action to destabilize a relatively hydrophobic ionic solute in aqueous solution through the formation of an ion-pair complex. Generally, a Lewis base reacts with a Lewis acid, e.g., Hg2+ binds S2. In some cases, a simple pH modification is enough. Hong et al. (2002a) were able to remove exchangeable and carbonated Cd(II) and Zn(II) from three different soils using Saponine at 3.7%; they recovered up to 78% surfactant in treated leachates after a metal precipitation stage at pH 10.7. Alternatively, dissociation in highly acidic medium was used for the EDTA recovery, before MTEs precipitation in alkaline medium (Di Palma et al. 2005; Pociecha and Lestan 2012). Despite the high recovery yields achieved, up to 88%, this strategy requires a prior concentration step and several pH adjustments that consume important quantities of inorganic reactants and increase the salinity of concentrates. MTEs have been displaced from chelants and then removed from soil leachates at pH 9 as oxyhydroxides in addition to As(V) using excess of Mg (Ehsan et al. 2007; Rivero-Huguet and Marshall 2011; Wen and Marshall 2011; Wen et al. 2012). It must be noticed that alkaline-earth metal chelates are still very reactive for metal extraction in soils, but the use of elementary alkaline-earth metals is too expensive for the regeneration of chelating agents. The low solubility of MTE– sulfide complexes was used to regenerate chelating agents (Hong et al. 1999, 2002b; Zeng et al. 2005), but the use of free sulfide ion as reagent is very dangerous. Efligenir et al. (2013) has shown that the regeneration of polyaminocarboxylate ligands in MTE-contaminated soil leachates was achieved within few hours of contact with FeS at pH 5. The resulting iron chelates were further transformed in the form of reactive Ca complexes in presence of calcium phosphate in order to remove iron as FePO4(s). Despite it being a two-step process, the low reagents cost, the high efficiencies of reactions, and its low sensitivity to iron-rich soils make it especially interesting. Diethyldithiocarbamate is also very effective to remove MTEs from chelates (Xie and Marshall 2001), but its high cost prevents its use from regenerating most of chelating agents. Coagulation–flocculation is a two-step process in which charged contaminants in wastewater are first neutralized, using polyions of opposite charge, before the size of the colloidal product obtained is increased, using a low solubility and high molecular weight polymer, in order to enable rapid particles settling. The removal of the chemical oxygen demand (COD) from the PAH-contaminated soil leachates using coagulation was studied for anionic, cationic, and nonionic surfactants (LopezVizcaino et al. 2012). Trivalent iron and aluminum salts were used as low-cost reagents to remove micelles and emulsions. As expected, the separation occurred efficiently for the anionic surfactant only. A COD removal larger than 90% was obtained for huge amounts of Al(III), but the remaining surfactant was not quantified. Ahmed Mohamed (2014) has shown that up to 94% of petroleum hydrocarbons

1 Contaminant Mobilization from Polluted Soils: Behavior and Reuse of Leaching. . .

39

in leachates (TPH approximately 1000 mg l1) from a creosote-contaminated soil could be removed while the SDS recovery was about 9% only. The main concerns with this method is the risk to favor anionic surfactant precipitation in soil leading to soil pore clogging when the coagulant is introduced in excess during the wastewater treatment. Cationic surfactant like cetyltrimethylammonium bromide (CTAB) have been proposed as an alternative to inorganic coagulants (Hirasaki et al. 2011), but their price precludes their use in this context. Electrocoagulation was proposed for the simultaneous regeneration of EDTA at pH 10 from Pb–EDTA complexes (Pociecha and Lestan 2010). Lead was recovered both at the cathode and in Al oxyhydroxides; however, Faraday yields for metal recovery are usually low in alkaline conditions (Awal Abdilahi 2015) and ligand recovery depends strongly on pH value. Similarly, Ferraro (2015) reports the use of iron electrodes for the regeneration of Cu-contaminated EDDS solutions. However, despite Cu being removed from EDDS, the new form of the ligand was unclear and some caution must be taken with respect to regeneration, since Fe complexes of EDDS are kinetically quite inert and have stability constants similar to Cu–EDDS.

1.4.2.3

Solvent Extraction

This technique is usually used for the removal of neutral contaminants or ion pairs obtained with a companion molecule. It has been used for the fast regeneration of cyclodextrins from PAH-contaminated soil leachates using colza oil (Petitgirard et al. 2009). A pilot-scale experiment using a closed loop automated system showed the preserved activity for the regenerated solution despite the numerous performed cycles, thanks to the quantitative CD recovery through regeneration. Despite the interest of this low cost, high efficiency, and eco-friendly regeneration process, the use of CDs has limited interest for field applications. The regeneration of HOC-contaminated surfactant solutions has been also tested with various organic solvents (Lee et al. 2002; Ehsan et al. 2007; Ahmed Mohamed 2014). However, there is a real concern for secondary pollution since their solubility in water is still enhanced by surfactants. The higher the surfactant concentration is in aqueous solution, the lower the partition constant is for HOCs between the organic phase and the aqueous phase (Kungsanant et al. 2008). For creosote-contaminated soil leachates with surfactant concentrations at 2%, the partition constant ranged from 80 to 4, following the order ethyl acetate > petroleum ether > diethyl ether > methyl tert-butyl ether (MTBE). The nature of the polar head of the surfactant has little influence on the value of its o/w partition coefficient. The use of organic solvent represents a real danger for workers and moreover, it changes the properties of the treated wastewater; hence, except for ex situ treatments, it is not recommended.

40

1.4.3

N. Fatin-Rouge

Chemical Treatments

This strategy should be used with caution since some difficulties may potentially occur. The main issues are related to both the use and the storage of chemicals on site, the potential increase in waste volume, the occurrence of hazardous or unwanted reactions. Unwanted reactions in the treated soils or within the regeneration units may hinder or even quench the expected reactions and lower treatment yields. For example, oxidation is rarely specific, while soil and leachates are loaded with various components that consume oxidant (Haselow et al. 2003). Finally, the use of chemicals in the regeneration units may also over-contaminate the treated sites if they are carried along in the reused leaching solutions.

1.4.3.1

Oxidation

The selective degradation of PAHs from soil leachates in presence of CDs or Tween 80 using electro-Fenton oxidation has been studied (Mousset et al. 2014, 2016; Trellu et al. 2017). PAHs degradation larger than 99% was achieved after 4–8 h, while the recovery rates for ligands were higher for CDs (approximately 90%) than for Tween 80 (50–79%), because of different mechanisms and accessibility to PAHs in the complexes. The best conditions for the selective degradation of PAHs in presence of Tween 80 were observed at low current density. Despite the toxicity of treated leachates remained high, in particular because of the production of more hydrophilic oxygenated metabolites, their reuse for soil washing did not show any loss for soil bioactivity. The reported energy consumption for this electro-Fenton process was in the range 0.36–0.69 kWh g1 TOC removed. Considering a usual TOC about 800 mg l1 for a PAH-contaminated soil leachate, the energy consumed for degradation would be higher than 90 € m3 of treated wastewater.

1.4.3.2

Reduction

A catalytic hydrodechlorination was used to partially degrade pentachlorophenol into cyclohexane. This soil leachate was obtained from poly-contaminated soil leachates (Rivero-Huguet and Marshall 2011) in the presence of EDDS and Brij 98 (nonionic surfactant) after solvent extraction. The reported reduction yield was 60% and whereas it did not seem of immediate economic interest, the idea of degrading organic contaminants by reductive ways is attractive and deserves to be mentioned.

1 Contaminant Mobilization from Polluted Soils: Behavior and Reuse of Leaching. . .

1.4.3.3

41

Transmetallation

This is an organometallic reaction that involves the transfer of a ligand from one metal to another. Lim (2005) reports the use of Fe(III) in excess to remove MTEs divalent cations from their EDTA complexes, before EDTA regeneration in presence of a large amount of phosphate in strong alkaline medium, in order to remove iron (III) as solid FePO4.

1.4.4

Conclusion

A brief review of technology assessments for the reuse of soil leaching solutions is presented in Table 1.2. Even if the regeneration of contaminated leaching solutions has been demonstrated, its interest is not currently obvious since most of the available treatment technologies suffer from selectivity, efficiency, and cost. The process combining contaminant extraction and wastewater treatment must be considered as a whole. Its optimization for efficiency and selectivity with respect to reactants, technologies, and process parameters is complex (Ahn et al. 2008b; Hanafiah et al. 2018). The proper technology for leaching solution recovery should follow the order of preference: physical > physicochemical > chemical, since risks for operators, the natural environment, process disruption as well as waste management and recycling should be considered. Also one should bear in mind that chemical reactions are never specific and that their efficiency and safety are limited by interferences. Moreover, the number of steps and tools involved in treatment trains should be lowered as much as possible in order to avoid complexity, fragility, and costs. Currently, the most attractive technologies are: • UF for the removal of HOCs, since it is robust (single step treatment) and economic (approximately 1.25 € m3 of treated wastewater), but it is limited by surfactant recovery from concentrated wastewater • Pervaporation for the removal of VOCs (approximately 4.7 € m3 of treated wastewater), but on diluted wastewater • Sorption onto chelating resins such as Amberlite IRC 748 for the removal of MTEs when leachable amounts of iron or nickel are low, since resins are easily regenerated in acidic medium and may be advantageously coupled to electrolysis for metal precipitation and purification • Precipitation with sulfide ions for the removal of MTEs, but it may be dangerous and the strategy for the soil treatment itself should be carefully designed. Finally, it is a pity that most of the reported trials have been performed at the lab scale, since field tests are required for validation and for full assessments. In the next section, a field test assessing the feasibility of using UF for the reuse of soil leachates in SF operations for the removal of petroleum hydrocarbons with PAHs as major components is described.

Physicochemical

Method Physical

Precipitation • Neutralization

Sorption

Ultrafiltration

Pervaporation

Treatment technology Air-stripping

S ~ 88–96 S ~ 80–90 ChAg 30–100

Phen  90

HCB 80–99 MTE  90

S nonionic 78 ChAg ~ 90 ChAg  78 ChAg

MTEs 80–90

MTEs > 90

MTEs  68

MTEs

PAHs 90–99 9

S anionic ~100 S nonionic S nonionic S nonionic 46–80 S 5–80

S nonionic 90 S anionic ~100

Extractant recovery yield (%) S anionic ~100

PCE 96 TCE/PCE 90–95 Toluene 10–60 BTEX > 90

Contaminant removal yield (%) Naph 20–60 TCE 30–80 BTEX 63–74 TCE 5–90

DTC

conc. H2SO4 S2 (Na, Ca) FeS

PAC resin

PAC

Reagent

Effective, fast

Low cost

Fast, effective

Low energy

Single step treatment High removal yield/125 € m3

Advantages High reactant recovery 4.7 € m3 of treated solution

Table 1.2 Brief review of treatment technologies assessments for the reuse of soil-leaching solutions

Too expensive

LTT/ LWV

Dangerous

Requires concentration

–

LTT

Low S recovery from concentrated leachates

Inconvenients Low removal yield at high Cs Need prefiltration May need reconcentration step Low flux at high Cs

Hong (1999, 2002b) Efligenir et al. (2013) Xie and Marshall (2001)

Hong (2002b)

Hanafiah et al. (2018) Ahn (2007, 2008a, 2010) Wan (2011) Ahmed Mohamed et al. (2013)

Sabatini (1998) Abou-Nemeh et al. (1999) Vane (2001) Kim (2007) Topf (2013) Sabatini (1998)

References Lipe (1996)

42 N. Fatin-Rouge

EDDS

Cu2+ ~ 100

PAH > 99

PCP 60

MTEs 90–95

Oxidation

Reduction

Transmetallation

ChAg 84

CDs 90 S 50–79 S nonionic

CD 90–95 S 55–70

CD ~ 100 S anionic

S anionic EDTA  88

TPH  90 Pb2+ ~ 68

Phen, pyr ~ 90 Toluene TCB 98 PCB PAH/PH 98

S anionic  10

PAH  94

Org solvents ElectroFenton Hexane, Pd, H2 Fe(III) NaOH Ca3(PO4)2

CTAB Alelectrod. Feelectrod. Colza oil Organic solvents

Al3+/Fe3+

0.69 kwh g1 TOC Selective degradation Low cost

Fast, effective

High flow rate

Ehsan (2007) Ahmed Mohamed (2014) Mousset (2014, 2016) Rivero-Huguet and Marshall (2011) Lim (2005)

– – Stable oxygenated metabolites For Ex Situ only Low degradation rate Multi-steps, LWV

Petitgirard (2009) Lee (2002)

Lopez-Vizcaino et al. (2012) Hirasaki (2011) Pociecha and Lestan (2010) Ferraro (2015)

Too specific High risks

Too expensive

Risk for S precipt.

Cs surfactant concentration, CDs cyclodextrins, ChAg chelating agent, CTAB cetyltrimethylammonium bromide, DTC diethyltlhiocarbamate, HCB hexachlorobenzene, LTT long treatment time, LWV large waste volume, MTEs metal trace elements, Naph naphthalene, PAC powdered activated carbon, PCP pentachlorophenol, phen phenanthrene, pyr pyrene, S surfactant, TCE trichloroethylene

Chemical

• Solvent extraction

• Coagulation/ floc.

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44

1.5

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Example of Field Assessment of Soil Flushing with Reused Solutions (SFWRS)

SF tests with reused fluids were performed in Solec Kujawski (Poland) at the EU-project Timbre test site with the aim of assessing technological efficiencies and costs (http://www.timbre-project.eu/news-volltext/items/timbre-in-soleckujawski.html). The site was a former wood impregnation industrial plant operated from 1876 to 2001. It had a surface of about 13 ha, while about 4 ha were heavily polluted. The GW level was lower than 3.5 m below the soil surface. The soil was sandy and the plot of land used for experiments was contaminated because of dripping from impregnated woods. A preliminary visit for site recognition and sampling was carried out. Sampling was performed in an open area while the SFWRS test was carried out closer to the main entrance of the site for practical reasons. Soil samples collected during the sampling campaign were used for soil treatability studies and to design the wastewater treatments for the reuse of soil leachates. The soil porosity and apparent density were 0.182 0.073 and 1.33, respectively. Its intrinsic water permeability was 4 D and its pH was 5.5. The details of soil and contamination characteristics obtained for the collected samples are reported in Table 1.3. The main targeted contaminants were PAHs and C10–40 total petroleum hydrocarbons index (TPH). Several anionic (SDS) and zwitterionic (cocamidopropylbetaine, CB and laurylbetaine, LB) surfactants were assessed for TPH extraction in soil leachates at CS ¼ 2%w, much larger than their CMC. Betaines-derivative surfactants were selected based on their biodegradability and low toxicity (Holmberg et al. 2002). In lab tests, surfactant foams were used to flush the contaminated soil in order to improve HOCs extraction during the treatment (Mulligan and Eftekhari 2003; Wang and Chen 2012). In the best case, flushing in soil columns at the lab scale required a minimum of 20–50 pore volumes (PV) of surfactant solution injected for the full extraction of petroleum hydrocarbons; three quarters of the extracted TPH were removed in the foam gas phase (Table 1.4). The loss of betaines surfactants in soils was much lower than for SDS (Fig. 1.13). Finally, LB was selected for field tests, because it produced low-persistent foams required for SF and cross-flow UF operations. The linear part for the exponential extraction rate corresponding to the most productive part has been observed up to 5 PV, which has been the targeted level for injections in field-scale tests. An in situ field test was carried out at the pilot scale by excavating a contaminated plot (dimensions: 2 m long, 2 m large, 3 m depth). A 2 mm-thick PET sheet was set at the bottom and folded vertically on sides (0.5 m high) to make the cell impervious. Then a central well was installed to collect soil leachates and the soil was set back and compacted using a mechanical shovel. Finally, the top soil was clad with a 2 mm-thick PET sheet folded down on sides (0.3 m) and 16 injection–extraction wells were set. Considering the dimensions of the cell and soil characteristics, the PV was estimated to be 2.4 m3. All the steps and details about the setup of the cell are shown in Fig. 1.14, while details for wastewater treatments are shown in Fig. 1.15.

1 Contaminant Mobilization from Polluted Soils: Behavior and Reuse of Leaching. . .

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Table 1.3 Measured characteristics of the contaminated soil calculated as the 95% confidence level of three analyses Soil Conductivity (mS cm1) 0.45 Particle size distribution 0.01 at 20  C) • Cannot be used for free product less than 1 cm in thickness • Longer initial start-up and adjustment periods than with other conventional LNAPL recovery techniques; specific skills are required for operating the unit • Applies to water tables less than 7 m deep • Costs are generally higher than other treatments, and highly dependent on treatment duration, water, and pumped airflow rates, and number of recovery wells • Radii of influence must be closely analyzed before treatment is implemented • Requires implementing both water and gas treatment units • Contamination may spread if groundwater flow direction is not controlled • High humidity in the unsaturated zone reduces air permeability, thus hindering extraction • Significant variations in water table depth are a major obstacle for this process to function properly

Compiled from USEPA (1996, 1997a, b, 1999), Suthersan (1997), Colombano et al. (2010)

Well diameter does not influence long-term DNAPL recovery, but allows highvolume pumping. Indeed, small diameter wells result in lower DNAPL–water mixtures. Wells equipped with strainers located upward from the portion affected by pure DNAPL will result in higher product recovery and limited water intake (Schmidtke et al. 1992). Different pumping techniques can be implemented with the aim to improve the efficiency of pure product recovery: upwelling, waterflooding, and trench systems (Connor et al. 1989). These different approaches must be optimized via feasibility/ treatability tests, and when possible, ad hoc modeling to secure the sites where remediation will take place.

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Fig. 2.16 Schematic representation of DNAPL free product recovery [adapted from Colombano et al. (2010)]

2.4.1

Free Product Recovery with Groundwater Extraction and Skimming (Upwelling)

The upwelling technique consists in pumping water above the DNAPL recovery zone. The cone of depression formed by pumping the groundwater will raise the DNAPL/water interface level (Villaume et al. 1983; Wisniewski et al. 1985; Ferry and Dougherty 1986; Cazaux et al. 2014). The process involves using specific hydrocarbon pumps and dedicated water pumps. These are placed either in the same recovery well or in two separate wells (Fig. 2.17). This technique can significantly increase the DNAPL recovery level in the wells. Recovery flow rates 2–3 times higher than with conventional pumping techniques have been reported for dimethyl phthalate pollution (Wisniewski et al. 1985). However, the groundwater pumping rate must be very carefully determined: too high, and the flow rate will produce emulsions; too low, and the flow rate will not produce any upwelling. As a first approach, this phenomenon can be estimated with the Ghyben–Herzberg equation, used to formulate saltwater intrusion in coastal areas (Eq. 2.39).

z¼

 ρf h ρs  ρf f

ð2:39Þ

where, .: thickness of freshwater zone below sea level to a point on the interface (m)

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Fig. 2.17 Schematic representation of free product recovery by groundwater extraction and skimming (upwelling) [adapted from Villaume et al. (1983)]

.f: thickness of freshwater zone above sea level (m) ρf: freshwater density (kg m3) ρs: saltwater density (kg m3)

2.4.2

Waterflooding

Waterflooding, or hydraulic displacement, or dual-phase extraction take into account the fact that, under certain conditions, hydraulic gradients can cause the DNAPL pool to migrate (Craig 1971; Willhite 1986; Gerhard et al. 2001; Alexandra et al. 2012). The process comprises specific oleophilic skimmers, and dedicated water pumps allocated to groundwater pumping (just above the DNAPL/water interface). These are placed either in the same recovery well or in two separate wells (Fig. 2.18). Two methods are implemented to obtain a first estimate of the necessary pumping characteristics. 1. A DNAPL pool located on an impermeable substratum has the following characteristics: (a) capillary pressure increases with the pool’s internal depth; (b) capillary pressures are identical at the edges of the pool. When a hydraulic gradient is applied everywhere in the pool, the capillary pressures are higher in the downstream zone than in the upstream zone of the pool. These differences can cause the DNAPL pool to migrate. According to Kueper and Gerhard (2014), the

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Fig. 2.18 Schematic representation of the principle of waterflooding [adapted from Colombano et al. (2010)]

DNAPL will migrate if the left-hand side of Eq. (2.40) is greater than the righthand side, as follows: P ðLÞ  Pc ð0Þ Δρ L sin α þ Δh > c ρw g ρw

ð2:40Þ

where, Δρ: difference between DNAPL density ρD and water density ρw (kg m3) α: dip of the bed below horizontal ( ) .: gravitational acceleration (m s2) Δh: difference in hydraulic head between the top and bottom of the pool (.(0)  . (L)) (m) .c(L): capillary pressure at the bottom of the pool (Pa) .c(0): capillary pressure at the top of the pool (Pa) 2. Another approach considers the fact that DNAPL mobilization occurs when the interfacial tension (IFT) between the wetting (water) and non-wetting (DNAPL) phases decreases. This decrease in IFT coupled with the change of viscosity in the non-wetting phase allows the capillary pressure (which draws the DNAPL in the pores) to be overcome (Pennell et al. 2014). Pennell et al. (1996) used a method to estimate DNAPL mobilization in the porous medium (Pennell et al. 1996). This method uses two types of numbers: the capillary number (Nca), and the Bond number (NB). The capillary number can be expressed as in Eq. (2.41):

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N ca ¼

vw μw γcosθ

ð2:41Þ

where, .w: Darcy’s velocity of the wetting phase (upward direction is considered positive) (m s1) μw: viscosity of the wetting phase (N s m2) γ: IFT between wetting and non-wetting phase (N m1) θ: contact angle of the two-phase system ( ) The Bond number, .B, is a function of gravitational forces and capillary pressure (Eq. 2.42): NB ¼

gkk rw Δρ γ cos θ

ð2:42Þ

where, Δρ: difference in density between wetting and non-wetting phase (kg m3) .: gravitational constant (N kg1) .: intrinsic water permeability of the porous media (m2) .rw: relative permeability of the wetting phase (–) If we combine the two numbers, one can obtain the total trapping number, .T (Eq. 2.43): NT ¼

qffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffi N 2ca þ 2N ca N B sin α þ N 2B

ð2:43Þ

where, α: angle between the system flow direction and the horizontal direction ( ) Pennell et al. (1996) ran several experiments using perchloroethylene (PCE) in columns (packed with quartz sand). The experiments showed that for the PCE–water system, .T is equal to 2  105; at this value, the PCE stored in the pores starts to move. When .T is over 1  104 almost all of the PCE is removed (Pennell et al. 1996). It should be noted that pressure differences, even small ones, can favor DNAPL migration (Kueper et al. 2008). Once the water is pumped and treated, it can then be injected back upstream of the treatment zone to increase the hydraulic gradients. Moreover, in some cases, waterflooding allows significant DNAPL recovery. For chlorinated compounds for example, Alexandra et al. (2012) demonstrated that the ganglia-to-pool ratio (i.e., reduction in the pool fraction) could vary from 0.1 to 0.3, or even 0.7, depending on the type of DNAPL, degree of heterogeneity, and the applied hydraulic gradient (Alexandra et al. 2012).

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Pumping free product too rapidly into a recovery well, can break the DNAPL continuum, thereby halting the migration and recovery of this free product. To resume recovery, the threshold pressure must be exceeded, which means that either new recovery wells must be installed, or higher hydraulic pressures imposed upstream. Once the water is pumped and treated, it can then be reinjected upstream of the treatment zone to increase the hydraulic gradients. This technique generates an increase in DNAPL dissolution: (1) by forming ganglia (with higher DNAPL/water contact surface than the initial surface); (2) by changing the phase equilibrium due to contact between DNAPL and uncontaminated water (Miller et al. 1990; Imhoff et al. 1993; Nambi and Powers 2003; Grant and Gerhard 2007a, b).

2.4.3

Trench Systems

Trenches, backfilled with gravel packs, have been used successfully to recover DNAPL. Pumping systems can be active (skimmers are combined with water pumping), or passive (hydrocarbon pumping only). This system is similar to those presented in Sect. 2.3.2 with the exception that the drains are placed at the bottom of the aquifer, within the DNAPL zone (Fig. 2.19). This system is best suited in shallow aquifers. In deeper aquifers, directional drilling or specific excavations with ad hoc support can be implemented. Groundwater pumping can increase DNAPL recovery levels (Sale and Kuhn 1988; Huling and Weaver 1996), and can be used for both upwelling and waterflooding.

2.5

Improving DNAPL Recovery

Free product pumping is completed with pump-and-treat operations. These are usually lengthy operations (e.g., often more than 30 years for chlorinated solvents) (Harkness and Konzuk 2014), and are not very effective in the long term due to slow release from the residual saturation, and therefore slow remediation rates (Mackay and Cherry 1989; Travis and Doty 1990; Berglund and Cvetkovic 1995; Pankow and Cherry 1996). For these reasons, enhanced technologies are frequently implemented to limit the long-term costs, and length of these operations (USEPA 2003; Williamson 2014). Various enhanced technologies have been designed to reduce the mass of contaminants. The ones used for chlorinated compound recovery are presented in Table 2.8. For other DNAPLs (i.e., mainly Polycyclic Aromatic Hydrocarbons, heavy chlorinated compounds, coal tars, and creosotes), the enhanced techniques used are almost identical, with the exception of in situ chemical reduction techniques, in situ air sparging, and in situ bioremediation, which are less suitable, considering the product characteristics.

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Fig. 2.19 Schematic representation of trench systems [adapted from Huling and Weaver (1991) and Sale and Kuhn (1988)]

A study conducted by McGuire et al. (2006) on 59 sites contaminated by chlorinated compounds in the United States, compared the treatment duration and remediation yield of certain groundwater treatment techniques. The examined techniques were in situ biological reduction, in situ thermal desorption, and surfactantenhanced washing. The average remediation yield values for these techniques were found to be respectively: 95%, 97%, and 95% (Fig. 2.20). In addition, the treatment durations were remarkably shorter than with the conventional pump-and-treat approach (2–21 months versus several decades). The performance of these enhanced extraction techniques (desorption and washing) is explained, in particular, by their ability to considerably reduce the quantity of pure products present in the pores (by reducing interfacial tension, desorption of

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Table 2.8 Functional role for commonly used remediation technologies in generalized sequential treatment strategy (Williamson 2014) Mobile DNAPL extraction technologies Hydraulic displacement In situ thermal treatment Surfactant-enhanced extraction

Source zone primary treatment technologies Excavation In situ thermal treatment Surfactant/co-solvent flushing

Source zone polishing technologies In situ chemical oxidation In situ bioremediation In situ chemical reduction

In situ chemical oxidation In situ bioremediation In situ chemical reduction Soil mixing with ZVI or other reagent In situ air sparging

In situ air sparging Natural attenuation

ZVI zero valent iron

Fig. 2.20 Rebound assessment at source depletion sites: concentration reduction from before treatment to immediately after treatment, and at end of data record (chlorinated compounds) (McGuire et al. 2006)

contaminants from the solid matrix, and reducing product viscosity). Indeed, releasing a source of pure product (mass flow) will depend not only on groundwater characteristics and the primary physical and chemical characteristics of DNAPL, but also on the characteristics of each source: (1) magnitudes (particularly in the DNAPL/water interface); (2) ganglia-to-pool mass (GTP) ratio; (3) how the pores are connected (permeability); and (4) residual saturations (Miller et al. 1990; Imhoff et al. 1993; Nambi and Powers 2003; Falta et al. 2005a, b; Grant and Gerhard 2007a, b; Carey and McBean 2010a, b; Alexandra et al. 2012).

96

2.5.1

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Effect of Temperature on DNAPL Recovery

Thermal treatments have been used effectively for DNAPL recovery. Considering the example of coal tar, increasing the temperature can remobilize the residual tar by decreasing the values of the following parameters: density, interfacial tension with water, contact angle with water on a solid medium, and viscosity (Huling and Weaver 1996; Heron et al. 1998; U.S. Army Corps of Engineers 2014). Specifically, coal tar viscosity is very sensitive to temperature and can vary by one to two orders of magnitude when the temperature is increased from 20  C to 70  C (Baker et al. 2006; Brown et al. 2006; Birak and Miller 2009; Philippe et al. 2017). A few cases of thermally enhanced coal tar pumping have been reported in scientific literature, as they relate to field testing (McLaren et al. 2009). The authors studied thermal enhancement as a sustainable alternative technique. The coal tars were heated to 30  C, which reduced their kinematic viscosity by almost one order of magnitude (100–10 cSt). Globally, 22 m3 of tar were recovered in 6 months of pumping, with a 30% reduction in costs compared to conventional pump-and-treat methods. Thermal enhancement can produce high recovery yields (90%), and can be particularly interesting to treat highly contaminated areas with low permeabilities where flushing is not a suitable remediation technique (Suchomel et al. 2014).

2.5.2

Effect of Surfactant Addition on DNAPL Recovery

The recovery mechanisms during surfactant flushing include two main stages: (1) decreasing interfacial tension (IFT) and increasing contaminant solubility (NAPLs); (2) mobilizing the residual contamination (Pennell et al. 2014). (1) Decreasing IFT and Increasing Contaminant Solubility At low concentrations, surfactant molecules will mainly accumulate at solid–liquid or liquid–liquid interfaces (NAPL/water interface in our case, where a pure phase exists). Surfactant molecules will gradually cover the NAPL/water interface as surfactant concentration increases. Increasing surfactant concentration will reduce IFT until all NAPL/water interfaces are covered. At this stage, the increase in surfactant concentrations will no longer reduce the IFT: the surfactant molecules will agglomerate (forming surfactant micelles), and will increase the solubility of the NAPL (present in the dissolved phase). This concentration is called the Critical Micelle Concentration (CMC) (Vishnyakov et al. 2013). The CMC of a surfactant depends on surfactant structure, system temperature, ionic strength, and whether or not organic additives are present in the solution (Laha et al. 2009). Surfactant performance depends on their working environmental conditions. For example, system temperature and salinity can influence surfactant effectiveness. When the temperature increases, the reaction between the hydrophilic component

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of a nonionic surfactant and water decreases. For ionic surfactants, the reaction between the hydrophilic component and water decreases when the salinity of the system increases (Rosen and Kunjappu 2012). In general, interfacial tension decreases with increasing temperature (e.g., 5.5  105 N m1  C1 for crude oil–water systems), and may be affected by pH, the addition of surfactants, and other substances in the solution (Schowalter 1979). The interfacial tension in the DNAPL–water system is directly related to the capillary pressure through the interface (Mercer and Cohen 1990). The interfacial tension in a DNAPL–water system varies between zero for completely miscible liquids, and 72  103 N m1 for absolutely immiscible liquids (72  103 N m1 is the water/air surface tension at 25  C) (Lyman et al. 1982). (2) Mobilizing Residual Contamination Surfactants are used to reduce the IFT in order to (a) displace maximum pure products (DNAPLs), and (b) decrease residual saturations. DNAPLs are displaced when the reduced IFT coupled with the change of non-wetting phase viscosity overcomes the capillary pressure. Therefore, it becomes necessary to choose the optimal surfactant concentration which will improve the recovery yield, and thus reduce residual saturation. Previous studies have shown that all tested surfactants have the effect of solubilizing and reducing IFT of the TCE–water and PCE–water systems. For TCE, Aerosol-MA-80 (5 wt%) significantly reduces the IFT of the TCE–water system as it falls from 35.2 dyn cm1 to 0.2 dyn cm1 (Dwarakanath et al. 1999). Tween 80 (5 wt%) reduces it from 35.2 dyn cm1 to 10.4 dyn cm1 (Suchomel et al. 2007). As for PCE, the Aerosol family of surfactants is very efficient for reducing the system IFT from 47.8 dyn cm1 to less than 0.01 dyn cm1 (Dwarakanath et al. 1999; Sabatini et al. 2000; Childs et al. 2004). Triton x-100 and Tween 80 are also effective (Taylor et al. 2001; Harendra and Vipulanandan 2011). Field experiments, described in specialized literature, reported recovery yields of pure chlorinated solvents ranging from 60 to 70% (Rao et al. 1997; Holzmer et al. 2000; Jawitz et al. 2000; Brooks et al. 2004; Soga et al. 2004), or sometimes more than 90% (Londergan et al. 2001; Abriola et al. 2005; Ramsburg et al. 2005; Pennell et al. 2014).

2.5.3

Using Surfactant Foam for DNAPL Recovery

Recent laboratory studies on surfactant foam technology for in situ removal of chlorinated DNAPLs have shown that this technique presents a promising line of research (Maire et al. 2015, 2016): • High foam stability for .surfactant at 0.05% was maintained despite presence of DNAPL • Strong foams (finely textured foams) resulted in more than 95% DNAPL recovery yield with surfactant consumption below 10 g kg1 of DNAPL recovered • No DNAPL fragmentation or enhanced dissolution (