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Sustainable Use and Management of Soil, Sediment and Water Resources15th International Conference | 20–24 May 2019 | Antwerp, Belgium

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AquaConSoilAntwerp 2019

Proceedings

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Table of Contents

1. Soil and water in the digital world

1b1 Data management and visualisation 5

STATISTICAL MODELLING OF GROUNDWATER CONTAMINATION MONITORING DATA USING GWSDAT: A COMPARISON OF SPATIAL AND SPATIOTEMPORAL METHODSM.I. Low, M. Bonte, L. Evers, A.W. Bowman & W.R. Jones

5

2. Advances in assessment of risk and monitoring of soil, sediment and water quality

2a1 Innovative and combined approaches for high resolution site characterization 7

ADVANCES IN IN-SITU AUTOMATOUS LNAPL AND WATER LEVEL MONITORING BY GUIDED WIRE RADAR: DETAILED ANALYSIS OF LNAPL BEHAVIOUR AND IMPROVED SITE UNDERSTANDINGDavid Holmes

7

ENHANCED DATA COLLECTION IN THE FIELD: ONSITE, HIGH RESOLUTION VOC ANALYSIS IN SOIL AND AMBIENT AIR DURING INVESTIGATION AND REMEDIATION OF A CHLORINATED SOLVENT CONTAMINATED SITEWim Vansina, Arnout Soumillion, Yves Van den Bossche, Pieter Buffel & Samuel Van Herreweghe

17

2a2 Passive sampling and mass flux measurements 23

INNOVATIVE CHARACTERIZATION TOOLS FOR DYNAMIC MONITORING OF GROUNDWATER AND DISSOLVED CONTAMINANTSAnthony Credoz, Nathalie Nief, Rémy Hédacq, Béatrice Ducastel, Laurent Cazes, Salvador Jordana, Martí Bayer-Raich, Trevor Osorno & John Frederick Devlin

23

POLLUTION FLUX MEASUREMENTS AS PART OF RISK ASSESSMENT FOR THE SPREADING OF GROUNDWATER POLLUTION: EXPERIENCES FROM PILOT PROJECTSBert Van Goidsenhoven, Kurt Bouckenooghe, Karel Biesemans & Isabelle T’Seyen

33

2c Soil-sediment-water interaction and system dynamics 39

ROLE OF PERCHED AQUIFER IN THE FATE OF THE CONTAMINANTS: CASE STUDYPetr Kozubek 39

2d2 Human health and environmental risk assessment: framework, tools and practice 44

POLLUTED SITES. FRENCH METHODOLOGY, EXAMPLE OF A BEST PRACTICECharline Darracq 44

3. Diffuse and emerging contaminants in the soil-sediment-water system3c Reactive interfaces: unsaturated zone, sediments and groundwater 54RESPONSE PROJECT: REACTIVE TRANSPORT MODELLING OF POINT SOURCE CONTAMINATION IN SOILS AND GROUNDWATERBertrand Leterme, Lana Geukens, Diederik Jacques, Vanessa M.A. Heyvaert, Marijke Huysmans, Cas Neyens, Erik Smolders, Dirk Springael, Bas van Wesemael & Jan Walstra

54

4. Advances in remediation technologies

4a1/2 Pilot and field scale biological reductive dechlorination 62

CHEESE WHEY INJECTION IN GROUNDWATER: USE OF AN ECONOMICALLY AND ECO-FRIENDLY SUBSTRATE FOR IN-SITU BIOREMEDIATION OF CHLORINATED SOLVENTSAndrea Campioni, Luca Sacilotto, Antonio Molinari, Giada Di Marco & Michele Leccese

62

ENHANCED REDUCTIVE DECHLORINATION OF PCE AND TCE IN A SOURCE ZONE VIA RECIRCULATION: PILOT TEST AND RESULTSPieterjan Waeyaert 64

4a 9/10 Advances in nanoremediation technologies 74

SYNTHESIS, CHARACTERIZATION AND REACTIVITY OF HYDROXYAPATITE COATINGS DEPOSITED ON CALCIUM CARBONATE POWDER FOR THE REMOVAL OF HEAVY METALS FROM CONTAMINATED WATEROriol Gibert, Vicenç Martí, Àlex Marín, César Valderrama, María Martínez & Rosa Maria Darbra

74

DEVELOPING ZERO-VALENT IRON NANOPARTICLES (NZVI) SUSPENSIONS FROM PLANT EXTRACTS AND ASSESSING THEIR REACTIVITY TO HEXAVALENT CHROMIUMMichalis Karavasilis & Christos D. Tsakiroglou

80

COMPARATIVE ADSORPTION OF GROUNDWATER CONTAMINANTS ONTO DIFFERENTS SIZES OF PARTICLES OBTAINED BY TWO TOP-DOWN APPROACHESVicenç Martí, Irene Jubany, David Ribas, José Antonio Benito, Berta Ferrer & Ada Ginesta

89

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AquaConSoil 2019 Proceedings

4a11 Electro-based (bio)remediation technologies 95

CATHODIC AUTOTROPHIC COMMUNITIES FOR STIMULATING Cr(VI) BIOELECTROCHEMICAL REDUCTIONGabriele Berettaa, Anna Espinozac , Giorgio Ferrarib, Andrea Franzettic, Andrea Mastorgioa, Sabrina Saponaroa & Elena Sezennaa

95

ELECTRO-KINETICALLY ENHANCED NZVI: EXPERIENCES FROM FRANCE AND SWITZERLANDVojtech Stejskal, Petr Kvapil, Jaroslav Nosek, Pierre Matz, Antoine Joubert, Cédric Palmier & Jaroslav Hrabal 105

4a General 112

FIXED-BED COLUMN STUDIES FOR THE BIOSORPTION OF METHYLENE BLUE ONTO BANANA PEELSA. Stavrinou, C.A. Aggelopoulos & C.D. Tsakiroglou 112

4b 1/2 Thermal treatment 118

SCALING UP OF DIOXIN CONTAMINATED SOIL THERMAL DESORPTION TREATMENT: LABORATORY TESTS AND PILOT CONCEPTION AT BIEN HOA AIRBASE, VIETNAMA. Jordens, H. Saadaoui & J. Haemers

118

THERMAL DESORPTION WITH SMART BURNERS – CASE STUDY FOR THE IN-SITU TREATMENT OF UNSATURATED SOILS AT THE REFINERY OF GELA (ITALY)Katia Pacella, Jan Haemers & Hatem Saadaoui

123

4b4 Challenges and complex issues 133

ELECTRO BIO RECLAMATION (EBR) AS PART OF A FULL SCALE SOIL REMEDIATION PROJECT IN GHENTJonas Wittocx, Art Lobs & Jeroen Van Acker 133

4b General 138

THERMAL CONDUCTIVE HEATING (TCH): COMPARISON OF POWER CONSUMPTION TO HEAT THE SOIL BETWEEN ELECTRICAL RESISTANCE AND GAS FIRED BURNERSJ. De Grunne, H. Saadaoui & J. Haemers

138

BETWEEN CARAVANS AND BOREHOLES – CHALLENGES OF PERFOMRING IN-SITU REMEDIATION ON A CAMPING SITE IN THE BLACK FOREST, GERMANYPetra Grill, Markus Kampschulte, Hans-Peter Koschitzky, Claus Haslauer & Oliver Trötschler

140

TOWARD THE DEVELOPMENT OF PHOTO-REACTORS FOR THE EFFICIENT TREATMENT OF WATER POLLUTED WITH DYESM.V. Karavasilis, C. Aggelopoulos, C.D. Tsakiroglou, A. Barnasas & P. Poulopoulos 150

IN-SITU METAL PRECIPITATION OF A NICKEL GROUNDWATER CONTAMINATION. RESULTS OF A COMPARATIVE LAB STUDY ON THE IMMOBILIZATION OF NICKEL USING A CARBON SOURCE AND A PERMANGANATE SOLUTIONDirk Paulus, Stijn Van Slycken, Nele Van de Maele, Jaap Steketee & Marian Langevoort

158

THERMALLY ENHANCED IN-SITU REMEDIAION USING STEAM-AIR INJECTION – BACKGROUND AND IMPROVEMENTS OF THE “DLI-DESIGN-TOOL” SOFTWARE FOR REMEDIAL DESIGNOliver Trötschler, Willis Awandu, Hans-Peter Koschitzky & Claus Haslauer

165

IN-SITU ZINC PRECIPITATION USING INORGANIC SULPHIDESDirk Van Look & Lars Van Passel 174

4c3 Combining chemical and biological remediation 181

FULL-SCALE APPLICATION OF EHC LIQUID TECHNOLOGY FOR THE ISCR AND ERD TREATMENT OF AN AQUIFER CONTAMINATED WITH TETRACHLOROMETHANE AND CHLOROFORMAlberto Leombruni, Linda Collina, Matteo Avogadri, Aldo Trezzi, Patrizia Trefiletti, Mauro Consonni & Caterina Di Carlo

181

4c General 190

SUCCESSFUL TRANSITION FROM EX SITU TO IN SITU REMEDIATION FOR CVOC-IMPACTED GROUNDWATERThibault Nodin, Jim Day, Mike Perlmutter & Mauricio L. V. Cavalin 190

4c Sps2/3 How to bridge the innovation gap 198

HOW TO BRIDGE THE INNOVATION GAP – FROM PILOTS AND PROTOTYPES TO PRICE LABELED PRODUCTSNiels Døssing Overheu, Hasse Milter, Per Loll & Julian Bosch 198

4d1 Phytoremediation and ecological engineering and nature based solutions 201

PHYTOREMEDIATION AND MYCOREMEDIATION OF CONTAMINATED LAND AT INDUSTRIAL SCALE: A COMPLETE TOOLKIT FOR NATURE-BASED SOLUTIONSOlivier Bastin, Yann Thomas & Gaylord Machinet

201

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4d Sps2 Constructed wetlands for cost-effective and energy-efficient remediation of plumes 209

OUR PRECIOUS GROUNDWATER GOES GREY… WHAT CAN WE DO?Nanne Hoekstra, Han Teunissen, Frank Pels & Paul Verhaagen 209

5. Strategies and management of contaminated land including legal, socialand economic aspects5aSps1 Legal day: International developments soil(remediation) policy 213

BARRIERS LIMITING THE REMEDIATION OF CONTAMINATED SITES IN ITALY AND POSSIBLE SOLUTIONSCarlo Collivignarelli, Roberto Bellini, Giorgio Bertanza, Cesare Bertocchi, Andrea Calubini, Angelantonio Capretti, Sergio Cavallari, Mara Chilosi, Nicola Di Nuzzo, Giovanni Faglia, Vincenzo Riganti, Alessia Savoldi, Carlo Scarpa, Maurizio Tira, Mentore Vaccari & Albino Zabbialini

213

5aSps3 Legal day: Soil, sediments and waste 217

HOW TO MANAGE BOTH COSTS, ACCEPTABLE RISK AND STAKEHOLDERS WHEN EXCAVATING MATERIAL CONTAINING ASBESTOS IN A CONTEXT OF BIG URBAN PROJECTS? Juliette Payet, Julien Matha & Thierry Gisbert

217

5b Remediation goals and strategies 222

GOVERNANCE STRATEGIES FOR FINALIZING CONTAMINATED AREASGeert Roovers & Ron Nap 222

ASSESSMENT OF KEY SUCCESS FACTORS OF A COMPLEX REMEDIATION BASED ON A CASE STUDY “THE MARIAKERKE ACID TAR DUMP SITE, BELGIUM”Joke Van de Steene, Tom Behets, Stefan Van den put, Carl De Cock & Werner Staes

228

5c Sustainable remediation 238

THE TRIAL OF SUSTAINABLE REMEDIATION BY SURF-JAPANYasuhide Furukawa, Testuo Yasutaka, Makoto Nakashima, Hisaichi Suzuki,Yuuta Natori 238

RELIEVING COMMUNITY STRESS RELATED TO SOIL REMEDIATION BY PERFORMING AN INNOVATIVE REAL-TIME ON-SITE AIR MONITORINGArnout Soumillion, Pieter Buffel & Samuel Van Herreweghe

240

5d Managing large scale industrial and agricultural pollution (water soil energy food nexus) 247

TOWARDS FINITE AFTERCARE: BIOREMEDIATION AND 3D MODELLING AT A FORMER MANUFACTURED GAS PLANT IN UTRECHT, THE NETHERLANDSTim Grotenhuis, Yanru Song, Nora Sutton, Suzanne Faber, Johan van Leeuwen & Jan Gerritse

247

7. Land, soil, water and sediment in the circular economy

7a2 Sustainable soil management 254

REUSE OF THERMALLY-CLEANED SOIL AND TARMAC GRANULATE: HEALTH, ENVIRONMENT AND PUBLIC EMOTIONSEllen Brand, Frank Swartjes, Arjen Wintersen, Michiel Rutgers & Piet Otte 254

SUSTAINABLE SOIL MANAGEMENT IN CENTRAL DENMARK REGIONArense Nordentoft & Susanne Arentoft 261

7b1 Reuse and upgrading of materials for improved ecological functioning 269

ON-SITE OR COMMERCIAL TOPSOIL MANUFACTURE USING EXCAVATED SOILS OR INERT MATERIALS: FULL-SCALE EXPERIENCE OF A PRE-EUROPEAN NETWORKOlivier Bastin, Yann Thomas, Lucas Demierbe & Gaylord Machinet

269

PHYTOEXTRACTION CAPACITY OF PHRAGMITES AUSTRALIS IN CONSTRUCTED WETLAND TO REDUCE POLLUTANTS FROM INDUSTRIAL WASTEWATERA. García-Valero, S. Martínez-Martínez, M.A. Terrero, A. Faz, M.A. Muñoz, M. Gómez-Garrido & J.A. Acosta

274

REUSE OF SEDIMENTS IN LANDSCAPE DEVELOPMENT: ECOTOXICOLOGICAL MONITORING OF A MOUND (PART OF THE VALSE PROJECT)Florian Liénard & Laurence Haouche

282

SOIL ENERGY AS SMART LOW CARBON TECHNOLOGY FOR COST-EFFECTIVE CLIMATE MITIGATIONNanne Hoekstra, Martin Bloemendal, Sara Picone, Marco Pellegrini, Alicia Andreu Gallego, Julián Rodriguez Comins6, Adrian Murrel, Tim Grotenhuis, Alessio Verrone, Hendrik-Jan Steeman & Sophie Moinier

292

ZUNUREC: PURIFICATION AND NUTRIENT RECUPERATIONJonas Wittocx, Art Lobs & Els Berckmoes 302

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Statistical modelling of groundwater contamination monitoring data using GWSDAT: A comparison

of spatial and spatiotemporal methods

M.I. Lowa, M.Bonteb, L.Eversa, A.W.Bowmana, W.R.Jonesc

a: University of Glasgow, School of Mathematics and Statistics, University Place, Glasgow G12 8QQ, United

Kingdom of Great Britain and Northern Ireland, Corresponding author: [email protected]

b: Shell Global Solutions International B.V., Lange Kleiweg 40, Rijswijk 2288GK, the Netherlands

c: Shell Global Solutions (UK), 40 Bank Street, Canary Wharf, London E14 5NR, United Kingdom of Great Britain

and Northern Ireland

Background and objectives

Field monitoring of groundwater contamination plumes is an important component of managing

risks for downgradient receptors and remedial strategies that rely on monitored natural attenuation.

Collection of groundwater quality data can however take a considerable effort and be associated

with high cost. Here, we investigated the relative merits of analyzing groundwater quality data using

spatiotemporal statistical modelling as implemented in GWSDAT compared to a spatial method (e.g.

kriging data for specific time steps) (Jones e.a., 2015)

We assessed the accuracy of both methods and implications for data collection requirements. The

aim of this work was to determine whether the quantity of data collected can be reduced, while

retaining the same level of estimation accuracy, by analyzing groundwater contamination data using

a spatiotemporal model which “borrows strength” across time, rather than a spatial model for

individual sampling events. This abstract provides a summarized version of a full paper on this

research published this year (McLean e.a., 2019).

Methods

For the comparison study the predictive performance of a spatiotemporal p-splines model was

compared with a spatial p-splines model and a spatial Kriging model. Kriging is currently one of the

most popular methods for spatial modelling and interpolation and thus was used in the study as a

reference to current practice. Spline-based models allow the complex relationships, often exhibited

in spatial and spatiotemporal data, to be estimated in a flexible manner by constructing the

estimated function with polynomial pieces (Evers e.a., 2015).

To capture the variability encountered under field conditions, we used three hypothetical

groundwater contamination plumes with increasing complexity developed with MODFLOW/MT3D.

The statistical models’ predictive performance was assessed by calculating the mean square

prediction error (MSPE) for 200 random well networks consisting of 45 wells. For each simulation,

observation data was removed in a step-wise manner, and at each step a statistical model was built.

MSPEs were calculated using the ‘true’ concentrations simulated for the three plume types and the

data of the well network.

Results

The results shown in Figure 1 that when all the data are used the spline models have a slightly

superior performance in comparison with Kriging. As data are removed, the performances of the

spatial methods deteriorate markedly in contrast with a much gentler rate of decline with the

spatiotemporal model. This demonstrates the benefit of ‘borrowing strength’ across space and time.

This particular example involves a substantial number of wells and observations, to allow the relative

performances of the spatial and spatiotemporal models to be demonstrated effectively. Clearly, the

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effectiveness of any model will be considerably reduced when the data become very sparse. The

results in Figure 1 also show, as expected, a slight decrease in performance of all methods as the

complexity of the plume increases.

Figure 1: Lower panels shows simulated plumes with increasing complexity. The upper panels show

the relationship between prediction error against for the % of data points removed. Predicted

surfaces from one simulation of the ‘observation removals' can be viewed in an R Shiny application

at: https://marnie-svst.shinyapps.io/gw-app/.

Conclusions

The simulation study presented demonstrates that a spatiotemporal model results in a smoother,

clearer and more accurate prediction through time compared to spatial modelling of individual time

steps. To achieve the equivalent accuracy with a spatial model, the network needs to be sampled

much more extensively. Leveraging the data obtained in this way can result in a reduced need for

drilling and sampling of monitoring wells, while better characterizing and managing environmental

risks. Freely available tools such as GWSDAT (http://www.api.org/GWSDAT) can facilitate

environmental professionals to use spatiotemporal models.

References

Evers, L., D.A. Molinari, A.W. Bowman, W.R. Jones en M.J. Spence (2015) Efficient and automatic methods for flexible regression on spatiotemporal data, with applications to groundwater monitoring; in: Environmetrics, vol 26, no 6, pag 431-441.

Jones, W.R., M.J. Spence en M. Bonte (2015) Analyzing Groundwater Quality Data and Contamination Plumes with GWSDAT; in: Groundwater, vol 53, no 4, pag 513-514.

McLean, M.I., L. Evers, A.W. Bowman, M. Bonte en W.R. Jones (2019) Statistical modelling of groundwater contamination monitoring data: A comparison of spatial and spatiotemporal methods; in: Science of The Total Environment, vol 652, pag 1339-1346.

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ADVANCES IN IN-SITU AUTOMATOUS LNAPL AND WATER LEVEL MONITORING BY GUIDED WIRE RADAR: DETAILED ANALYSIS OF LNAPL BEHAVIOUR AND IMPROVED SITE

UNDERSTANDING.

Session ThS 2a

David Holmes, Ecologia Environmental Solutions Limited. Building 711 & 712, Kent Science Park, Sittingbourne, Kent, UK, ME9 8BZ, +44 (0)1795 471611

[email protected]

Keywords: LNAPL, Baildown test, recoverability, saturation, site investigation, in-situ

Summary

Light Non-Aqueous Phase Liquid (LNAPL) contamination is a widespread issue. Understanding its mobility when released into the environment is key for mitigation and managing its impact. A baildown test can be used, but obtaining good quality data is difficult and especially on active sites can put personnel at risk.

Ecologia have developed a greatly improved way of monitoring LNAPL. One that is reliable, repeatable and above all, safe. A guided wire radar system is used to provide measurements of LNAPL that exceed interface probe data, both in terms of accuracy and repeatability. Time on site and contact with the LNAPL are both reduced, thereby reducing potential risks to personnel.

The system can be deployed on any site – options include remote access and ATEX rating. Ecologia’s team of qualified scientists can also put your data into context with a fully referenced and scientifically robust report.

1. Introduction

1.1 LNAPL

Groundwater contamination by Light Non-Aqueous Phase Liquids (LNAPL) is a serious and ongoing concern at many contaminated sites across the world (Feenstra et al., 1991). LNAPLs are typically multi-component mixtures of hydrocarbons that are lighter than water and are typically found in the vadose zone and upper portions of the groundwater. LNAPL is a functional description and does not refer specifically to any one kind of liquid; only that it exists as a separate phase.

LNAPL poses risks to the wider environment both through mass transfer and mass transport mechanisms. Mass transfer processes – dissolution and volatilisation – can give rise to impacts some distance from the source. The body of LNAPL may continue to act as a source of dissolved phase contamination for many years (Essaid et al., 2015).

Therefore, in most cases there is a need to remove the LNAPL. The most cost-effective methods for LNAPL recovery which is clearly mobile is by hydraulic pumping. However, as the saturation of the LNAPL is reduced in the soil, the mobility of the LNAPL also falls to a level where hydraulic recovery is ineffective and unsustainable. Figure 1 shows an idealised decline curve:

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Figure 1. An idealised curve showing falling LNAPL recovery rates over the duration of an LNAPL project.

LNAPL mobility is becoming more accepted as a remediation metric (Simon, 2014; Kimball, 2017), and so understanding the parameters that control it is important. However, while measuring the fluid characteristics, comprising density, viscosity and surface- and interfacial tension generally are accepted and robust methods, measurement of representative soil characteristics to derive van Genuchten parameters, for example, pose practical difficulties. Sale, (2001), outlined the key practical difficulties for LNAPL impacted sites in particular: unconsolidated sediments pose challenges in preservation, and LNAPL can readily drain from extracted cores, rendering any attempts to measure the in-situ saturation moot. There are also issues concerning the applicability of laboratory-measured hydraulic functions to conditions in the field (Wessolek et al., 1994). Several types of LNAPL may be present over a range of geologies at any given site, adding a complicating factor.

Therefore, transmissivity as a bulk measurement of LNAPL mobility is more useful.

1.2 Measurement of LNAPL Mobility

When LNAPL mobility is expressed as a transmissivity, the LNAPL type, the soil type, and the range of saturations and therefore, permeabilities of LNAPL in the soil are considered. It is a conceptually simple, bulk measurement of LNAPL mobility. It has use for site investigation work to assess the potential for LNAPL recovery, during remediation to assess progress, and at the end of a given technology’s use to demonstrate the mobility and migration risk from an LNAPL plume. As such, LNAPL transmissivity is being increasingly used as a performance metric across the industry and is gaining regulatory acceptance (Simon, 2012; Kimball, 2017).

The most widely reported method to measure transmissivity is a baildown test. Conceptually, this is quite simple: the LNAPL in a borehole is quickly removed, and the LNAPL recharge is measured. Internationally used standards (ASTM E2856 – 13) have been published to provide detailed step-by-step guidance. This test not only provides a direct measurement of the mobility, but also provides important information about the geological formation: whether the LNAPL is unconfined, confined or perched (Kirkman et al., 2013).

To carry out a baildown test, the depths to the top of the LNAPL and the top of the water are measured and the volume of the well and filter pack is calculated. This volume is then pumped out ‘instantaneously’ (defined as less than 1% of the total test time (ASTM E2856 – 13) and the LNAPL and water levels are tracked as the LNAPL recharges.

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The pressure gradient which induces LNAPL flow into the well is the drawdown. For a given well at a specific time, a greater drawdown will mean a greater inflow of LNAPL into the well. As the test proceeds, the drawdown returns to zero and no further LNAPL flows into the well. It is the change in drawdown over time in combination with the discharge of LNAPL into the well that is used to calculate the transmissivity.

A schematic of the baildown test process is provided in Figure 2, below.

Figure 2. An LNAPL-bearing well at equilibrium (left) and immediately after a baildown test (right). LNAPL is depicted in red, with more intense shades showing higher saturations, and groundwater is depicted as blue. The top dashed line shows the top of the LNAPL, the solid dark blue line the potentiometric head, and the light, dashed blue line the base of the LNAPL. Following removal of the LNAPL, the top of the LNAPL falls only a small distance. The groundwater is able to rise further up the well: the displacing pressure provided by the LNAPL has been removed.

The drawdown is limited by the potentiometric head of the groundwater. The LNAPL displaces the groundwater in the well and the potentiometric head represents the position of the groundwater if LNAPL were not present. When the LNAPL is purged during a baildown test, the displacement pressure is removed, and the water rises. Water has superior transport characteristics to the LNAPL: water is denser, has lower viscosity, is typically the wetting fluid and is present at higher pore saturations than the LNAPL.

1.3 The problem

LNAPL and water levels are virtually always measured using an interface probe. The interface probe typically operates through detecting the change in light attenuation between air and LNAPL, and the change in electrical conductivity between LNAPL and water.

In a controlled environment, an interface probe can reasonably detect 1 mm changes in the LNAPL thickness. However, field-monitoring of LNAPL is commonly less controlled, and measurement error is potentially much greater. Crucially, the measurement error can approach the measurement range in many baildown tests. This is recognised in the ASTM E2856 – 13, where LNAPL gauging during baildown testing is recommended to allow at least a 1.5 cm change in LNAPL thickness to avoid noisy data.

Figure 3 shows LNAPL drawdown for LNAPL with different densities and thicknesses in unconfined conditions. A less dense LNAPL will displace less water, so more of the in-well thickness will be above the potentiometric head and will provide a larger measurement range. A denser LNAPL will displace

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more water, so more of its vertical extent will be below the potentiometric head, and the measurement range will be smaller.

Figure 3. Change in maximum drawdown with LNAPL thickness for LNAPLs of different densities (Modified from Hawthorne, 2015).

Figure 3 shows that for an LNAPL with a density of 0.8 g/mL and a thickness of 100 cm, the maximum drawdown of an LNAPL is 20 cm. It is common however to find LNAPLs that have densities of 0.9 g/mL and above due to degradation, dissolution and volatilisation of the smaller components. LNAPL thickness are typically also less than 100 cm. Batu (2013) reported that LNAPL thicknesses are typically between 5 and 50 cm for a 0.9 g/mL LNAPL with a thickness of 50 cm, the maximum drawdown is 5 cm in unconfined conditions. Considering the filter pack recharge may then further reduce this to 2.5 cm or even further. Thus, with smaller thickness and denser LNAPLs the interface probe cannot be reliably used to measure the LNAPLs behaviour.

More viscous LNAPLs create even greater practical difficulties. When the LNAPL is viscous, it can coat the tip of the interface probe. This can lead to misleading readings both above and below the LNAPL. When the probe is lowered into the water sometimes the hydrostatic pressure will force the LNAPL off the tip, allowing the LNAPL: water interface to be read. However, this is not always the case, especially if the LNAPL is close to the density of water or especially viscous. Sediment in the wells can also affect the reading. To measure the air: LNAPL interface of a viscous LNAPL, the operator has one chance to register the top of the LNAPL before an optical probe needs to be cleaned. This can result in long periods of time and effort being spent to monitor the thickness of LNAPL in a well.

Issues also arise due to the transmissivity and recovery rates of LNAPL being strongly affected by the groundwater level and the position of LNAPL in the well (Becket & Huntley, 2015; Hawkins, 2013). A higher water level may occlude LNAPL, lowering the transmissivity. Furthermore, as groundwater fluctuates, the LNAPL may in some instances be brought into continuity with strata of different permeability, leading to dramatic changes in the LNAPL transmissivity. This can result in a condition where LNAPL can be recovered in significant quantities, but only when the groundwater is at a given level. Conditions may also change between unconfined, confined or perched.

Neither longer-term tests of LNAPL mobility or long-term monitoring to uncover geological conditions can provide as much detail as a baildown test. While decline curve analysis and tracer testing can provide a bulk assessment of the LNAPL mobility over time, a detailed understanding of the interactions between the LNAPL levels and mobility cannot be made.

0

5

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0 50 100 150 200

Dra

wdo

wn

(cm

)

LNAPL thickness (cm)

0.8 (g/mL)

0.9 (g/mL)

0.95 (g/mL)

0.98 (g/mL)

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A baildown test is therefore a useful and important activity. It allows the development of a detailed site conceptual model: providing information about the geological conditions and the mobility of the LNAPL.

The advance in LNAPL monitoring presented below aims to fulfil the need for accurate, repeatable and near continuous data. An important ancillary benefit is that it also offers improvements in site operator safety: operator hours are reduced, lowering exposure to active and/or construction sites. Data on the health and safety aspects of contaminated land work are sparse and so the improvements gained by minimising time on site and contact with LNAPL cannot readily be quantified. The presented advance does however follow recommendations for health and safety improvements in that contact with the hazard is reduced and consequently, the need for additional PPE is also reduced (Haslam et al., 2005).

2. Method

A Vega Vegaflex 81 guided wire radar sensor was used to simultaneously track two liquid interfaces. This is a device that uses a weak microwave signal in combination with time domain reflectometry to calculate the thickness of a low-dielectric liquid. LNAPLs, due to their low-polarity, exhibit low dielectric constants: water, being highly polar, has a higher dielectric constant. The microwave signal is weakly reflected by the LNAPL, but more fully reflected by the water. The return signal is then decoded by the sensor and a reading provided by the software. A schematic is presented in Figure 4, below. The figure uses the same scheme as Figure 2. A number of modifications were made to the sensor housing to allow it to function on contaminated land sites. The internal hardware, firmware or software of the Vegaflex were not modified in any way.

Figure 4. Schematic describing the operation of a guided-wire radar system in an LNAPL-bearing well. At ‘1’, a weak electromagnetic signal is sent by the sensor down a guide wire. A small fraction of this signal is reflected by the increase in dielectric constant between air and LNAPL (‘2’), and a large fraction of the signal is reflected by the great increase in dielectric constant from LNAPL to water (‘3’).

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3. Results

The system was tested in the laboratory. Vegetable oil was used as a model LNAPL in the first instance, and then repeated with diesel. The LNAPL probe was then used in several field applications to measure LNAPL mobility which are detailed below.

The LNAPL probe was used on a site where up to 0.6 m of LNAPL was reported in highly permeable gravels. The probe was set up to monitor the levels during a baildown test. These are shown in Figure5.

Figure 5. Fluid levels and LNAPL thickness during a baildown test with 0.6 m of diesel in gravels. Time 0 is equal to the start of the test. The probe sits slightly above ground, so fluid depths are given as m below datum.

The LNAPL and water levels during this baildown test behaved in a previously described manner (Huntley 2000): as LNAPL is removed, the LNAPL level and thickness falls, and in response to less displacement pressure, the water level rises. During recharge, the LNAPL level increases and water continues to recharge into the well, but at a slower rate than the LNAPL, and consequently, the LNAPL thickness increases. The drawdown returns to equilibrium conditions and LNAPL discharge (LNAPL flow into the well) is positive. These data allowed a rapid analysis of transmissivity. Re-arranging the data to show filter pack effects was not required: filter pack recharge could be readily identified in the fluid levels and are excluded from the analysis.

High viscosity LNAPLs pose a particular problem for baildown tests. A fluid with a high viscosity flows less readily than low viscosity fluids, and so the mobility and recoverability could reasonably be anticipated to be lower. However, made ground can be composed of gravelly material, which may have a hydraulic conductivity of up to 5000 m/d (Lewis et al., 2006). In such material, even lubrication and hydraulic oils can have high mobility.

A site investigation was carried out at a manufacturing site in the UK, where ongoing small releases of hydraulic oil had built up over several years and resulted in an LNAPL plume within the underlying made ground soils.

The saturation of this LNAPL had developed to the extent that the plume had begun to migrate and impact third party land.

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A sample of the LNAPL was obtained and the viscosity measured over a range of temperatures. The groundwater temperature was recorded as 13°C to 14°C, giving the LNAPL in the soil a viscosity of 380 to 400 mm2/s.

A baildown test was carried out on site, using the LNAPL probe to track fluid levels. The thickness of the LNAPL was 30 cm, and so the drawdown was correspondingly small. The data from this test were re-arranged to show the change in drawdown over time.

The baildown data showed that the recharge of the LNAPL was quite rapid, and equilibrium fluid levels were reached in around 20 minutes. Figure 6 demonstrates that it is possible to monitor previously unmeasurable LNAPL, with a limited drawdown.

Figure 6. Drawdown over time for the highly viscous LNAPL. These data show that pressure difference in the LNAPL between the formation and well returned to zero in around 20 minutes. The first three data points show filter pack recharge.

A period of recovery then took place, followed by a surfactant flush, using Ivey Sol 106. The baildown test was repeated, again using the LNAPL probe. The data were again re-arranged to show drawdown over time. The thickness at the time of the second test had increased to over 2 m, so the data were normalised to show comparison with the first test.

This is shown in Figure 7, together with the data from the first test, prior to LNAPL recovery.

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Figure 7. Comparison of drawdown over time, before and after recovery for a highly viscous LNAPL.

These data demonstrate that the LNAPL probe is capable of measuring viscous LNAPL at a range of thicknesses and mobilities, and can clearly show a reduction in mobility following recovery.

A baildown test, carried out in diesel-impacted soils, reported to be silty gravels, showed unusual LNAPL behaviour. The thickness was around 0.3 m, indicating that there may be moderate mobility. In this test, it was found that the water mobility was far greater than that of the LNAPL. The baildown test data are shown in Figure 8.

Figure 8. Baildown test data in a 0.3 m thick LNAPL. The water moves into the well at a much greater rate than the LNAPL.

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Figure 8 shows that following LNAPL removal, the LNAPL level and water levels both rise, which is consistent with previous descriptions. However, in this case, the water level rises faster that the LNAPL, creating an apparent negative LNAPL discharge into the well. To the author’s knowledge, this behaviour has not previously been reported in the scientific literature.

4. Conclusions.

On a very straightforward level, the data gathered using the LNAPL probe are more accurate and more precise than data gathered using an interface probe. The system is not limited by the need for a site operator presence and can operate indefinitely. As such, this system shows a significant leap forward in terms of data gathering.

More importantly, through capturing the high-quality data, it can allow the calibration and development of new models to quantify the volumes of LNAPL present, which can lead to improved site understanding and more informed remediation decisions. LNAPL theoretical understanding across the contaminated land industry has progressed a great deal in the last few years, following the publication of the first Interstate and Technical Regulatory Council (ITRC) guidance in 2009, the Contaminated Land: Applications in Real Environments (CL: AiRE) guidance in 2014, and a greatly updated version of the ITRC LNAPL guidance in 2018. However, the provision of adequate data from site has presented practical difficulties (SoBRA 2018) and has perhaps held back robust decision making. Ecologia’s guided wire radar system presented has demonstrated that high-quality data concerning the mobility of LNAPL can be captured, analysed and used as both a leading and ending metric for LNAPL recovery projects, and shed light on previously unreported LNAPL behaviour following a baildown test.

There are no data on the health and safety aspects of LNAPL works, and contaminated land as a whole has not been closely examined in terms of occupational health. However, recommendations do exist that emphasize the reduction of risk through improved design. By removing the need for a continuous site operative presence and reducing the contact between personnel and the LNAPL, the LNAPL probe fulfils this recommendation.

Potential further work for the LNAPL probe includes the use of longer-term monitoring in both field and laboratory conditions to understand the relationship between soil characteristics, LNAPL type and saturation, and the LNAPL’s response to groundwater fluctuations. Potential changes in the dielectric properties of LNAPL with composition change may also provide a line of evidence as to any chemical processes taking place in the LNAPL.

Crucially, it offers a safer and higher resolution method of monitoring LNAPL, which significantly contributes towards being able to understand and predict LNAPL behaviour in real environments.

References

ASTM E2856-13, Standard Guide for Estimation of LNAPL Transmissivity, ASTM International, West Conshohocken, PA, 2013,

Batu, V. 2013. Author’s reply. In Charbeneau, R., Kirkman, A. and Adamski, M., 2013. An assessment of the Huntley (2000) baildown test data analysis method. Groundwater, 51(5), pp.657-659.

Beckett, G.D. and Huntley, D., 2015. LNAPL transmissivity: a twisted parameter. Groundwater Monitoring & Remediation, 35(3), pp.20-24.

Essaid, H.I., Bekins, B.A. and Cozzarelli, I.M., 2015. Organic contaminant transport and fate in the subsurface: Evolution of knowledge and understanding. Water Resources Research, 51(7), pp.4861-4902.

Feenstra, S., Mackay, D.M. and Cherry, J.A., 1991. A method for assessing residual NAPL based on organic chemical concentrations in soil samples. Groundwater Monitoring & Remediation, 11(2), pp.128-136.

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Haslam, R.A., Hide, S.A., Gibb, A.G., Gyi, D.E., Pavitt, T., Atkinson, S. and Duff, A.R., 2005. Contributing factors in construction accidents. Applied ergonomics, 36(4), pp.401-415.

Hawkins, A.M., 2013. Processes controlling the behavior of LNAPLs at groundwater surface water interfaces (Masters Thesis, Colorado State University. Libraries).

Hawthorne, J.M. 2015. Selection of an Optimal Site-Specific Method for the Measurement of LNAPL Transmissivity. International Petroleum Environmental Conference, 22nd Conference Presentation.

Huntley, D., 2000. Analytic determination of hydrocarbon transmissivity from baildown tests. Groundwater, 38(1), pp.46-52.

Kimball, C.G., Hawthorne, J.M., Menatti, J.A., Rousseau, M. 2017. Light Non-Aqueous Phase Liquid Transmissivity Acceptance and Use by the Regulatory Community.. IPSO conference at Tulsa.

Kirkman, A.J., Adamski, M. and Hawthorne, J.M., 2013. Identification and assessment of confined and perched LNAPL conditions. Groundwater Monitoring & Remediation, 33(1), pp.75-86.

Lewis, M.A., Cheney, C.S. and O Dochartaigh, B.E., 2006. Guide to permeability indices.

Sale, T., 2001. Methods for determining inputs to environmental petroleum hydrocarbon mobility and recovery models. American Petroleum Institute, Public, 4.

Simon, J.A., 2012. Editor's perspective—LNAPL transmissivity: Mobility is the new state-of-the-art clean-up metric. Remediation Journal, 22(3), pp.1-8.

SoBRA (Society of Brownfield Risk Assessment) Conference: Fine Tuning DQRA’s for the Water Environment. Workshop 4: When to use the API calculator, the input parameters and key lessons. 18th June 2018, held at Camden Town Hall, London, WC1 HJE.

Wessolek, G., Plagge, R., Leij, F.J. and Van Genuchten, M.T., 1994. Analysing problems in describing field and laboratory measured soil hydraulic properties. Geoderma, 64(1-2), pp.93-110.

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Poster session – 2a. Advances in sampling, monitoring techniques and methods

ENHANCED DATA COLLECTION IN THE FIELD: ONSITE, HIGH RESOLUTION VOC ANALYSIS IN SOIL AND AMBIENT AIR DURING INVESTIGATION AND REMEDIATION OF A CHLORINATED SOLVENT CONTAMINATED SITE

Wim Vansina, Wittenveen+Bos, Gorislaan 49, 1820 Steenokkerzeel, Belgium, Tel. +32 (0)2 759 59 30, [email protected] Arnout Soumillion, Wittenveen+Bos, Gorislaan 49, 1820 Steenokkerzeel, Belgium, Tel. +32 (0)2 759 59 30, [email protected] Yves Van den Bossche, OVAM, Stationsstraat 110, 2800 Mechelen, Belgium Pieter Buffel, EnISSA, Gorislaan 49, 1820 Steenokkerzeel, Belgium, Tel. +32 (0)2 759 59 30, [email protected] Samuel Van Herreweghe, EnISSA, Gorislaan 49, 1820 Steenokkerzeel, Belgium, Tel. +32 (0)2 759 59 30, [email protected]

Introduction

From the 1950’s up to 2001 a wool factory was located on a site situated in the center of a small village - Dikkelvenne - close to Ghent, Belgium. Those former industrial activities have led to a contamination of soil and groundwater with chlorinated solvents. Because of the many natural water sources welling in the vicinity of the village center, Dikkelvenne is known in this region as “the village of springs”.

Figure 1 Natural water spring

Recently chlorinated solvents have been found in some of the sources about 150 m south of the site and local residents are advised to stop using this spring water. OVAM (the public waste agency of Flanders) has appointed Witteveen+Bos Belgium to design a remediation concept for this contaminated site.

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High Resolution Site Characterisation with EnISSA-MIP and HPT

The success of soil investigations and remediation designs is highly depended on a solid and constantly adjusted conceptual site model (CSM). Underestimation of the contaminated area and or pollutant load can lead to an inadequate remediation design resulting in project failure, budget overshooting and residual risks. A detailed visualization of the contaminant situation will provide the consultant an improved insight in the source, pathways, exposure and remediation alternatives but will also help to recognize and identify the many data gaps and uncertainties. The CSM is built by collecting data from the site under investigation. The whole process of delineation and mapping of the contamination can become a very intensive and time consuming task. Traditional sampling methods are characterized by low detection levels and a broad and accurate analysis spectrum but despite the high accuracy and precision of the analysis at the certified labs, a large portion of the accuracy is already lost during the sampling, handling and conservation in the field. Environmental consultants are sometimes blindfolded by the final analytical certificate which give them a false sense of certainty. Furthermore the number of data points (samples) is often very low compared to the investigated area and volume. Since hydrogeology and contaminant concentration can vary on a small scale (centimeters), a limited data set can be problematic and will result in a CSM with considerable data gaps, uncertainties and wrong conclusions.

To update and verify the contaminant delineation and conceptual site model described in the previous investigation stages of the site under investigation in Dikkelvenne, Witteveen+Bos Belgium, decided to carry out a number of EnISSA-MIP probing. This GC-MS based improved version of the Membrane Interface Probe system, generates detailed depth profiles of the VOC’s. Because of the sensitive detector also the delineation of plume concentrations (10’s of µg/L) for the different degradation

Figure 2 Groundwater contamination with chlorinated solvents

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products is possible. The acclivitous character of the surface in this project area is an important additional difficulty to understand the contaminants migration and full depth profiles are even more important to avoid missing the contaminant plume.

Figure 3 Example of an EnISSA-MIP profile for PCE, TCE, DCE and VC

The EnISSA-MIP investigation has been combined with the Hydraulic Profiling Tool to collect more data on soil permeability . This extension of the MIP uses a water screen and pump to inject a constant flow of water through the probe. The adjusted pressure to maintain this flow is a measure for the local permeability of the soil. This tool allows for a detailed investigation of the different migration pathways through the more permeable layers and the soil layers where to expect ‘storage’ of the chlorinated solvents. High permeability intervals were detected and correlated with the presence of sandstone and boulder deposits.

Figure 4 Example of a HPT profile

However the EnISSA MIP investigation was only used in a later stage of the investigation on this site and the number was rather limited, the 7 EnISSA-MIP logs performed to check and update the CSM resulted in 210 data points for each evaluated compound, or 70 ’information meters’. A comparable investigation effort, i.e. 7 monitoring well’s with 1 or even 2 screens would have resulted in only 7 or 14 data points or ‘information meters’. A 3D application was used to visualize the results.

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Figure 5 3D visualization of the results

On-site GCMS analysis of CVOC’s in ambient air

The first phase of the remediation of the site has started recently with the removal of the source zone by excavation. Due to the situation in the middle of a residential area and the close proximity of a school, a close monitoring of the ambient air was requested. Due to the transport, typical slow lab processing and limited ‘multiple-hour-averaged samples’, a common approach to measure CVOC’s based on sampling with activated carbon tubes would lack speed and data density and intervention or adjustment of procedures based on instantaneous risks is not possible. PID detection is routinely applied to monitor excavation works in remediation projects. The small handheld detectors are very practical and have fast response times but results cannot discern individual compounds which makes comparison of the results to standard values that are individually defined not possible. Alternatively Witteveen+Bos Belgium suggested the application of EnISSA’s on site GCMS system to collect more data.

Figure 6 On-site GCMS system during excavation

The fast GCMS application is used to sample the ambient air and results for the selected CVOC compounds are available almost every minute. An automated data interpretation script is programmed to send a message to the on-site environmental supervisor once the air standard for each different parameter (in this case PCE, TCE, cis1-2DCE or VC) is been exceeded.

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Figure 7 Results of PCE, TCE, cis1-2DCE and VC in ambient air during remediation works

Results for this project showed a clear link with excavation activities during the day and a compound distribution as expected by the soil contamination: i.e. concentrations of PCE were the highest. However due to the more stringent standard for VC, the detected lower VC concentrations seemed to be critical as well. The increased level of detail of VOC’s in ambient air and ability to evaluate individual parameters is essential for a proper monitoring and necessary to have the ability to intervene the excavation actions when risks arise for the neighboring citizens.

Figure 8 Concentrations of PCE in ambient air during remediation works

Figure 9 Concentrations of VC in ambient air during remediation works

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Conclusion

The success of soil investigations and remediation designs is highly depended on a solid and constantly adjusted conceptual site model (CSM). EnISSA-MIP probings generates detailed depth profiles of the individual VOC’s. The combination with the Hydraulic Profiling Tool allows a detailed investigation of the different migration pathways of the chlorinated solvents

Due to the low detection limits and the possibility to measure individual components, an on site fast GCMS system that was used to sample the ambient air makes it possible to intervene the excavation actions when risks arise for the neighboring citizens.

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Oral presentation, Aquaconsoil 2019, session ThS 2a2 Passive sampling and mass flux measurements

INNOVATIVE CHARACTERIZATION TOOLS FOR DYNAMIC MONITORING OF GROUNDWATER AND DISSOLVED CONTAMINANTS

Anthony Credoz1*, Nathalie Nief1, Rémy Hédacq1, Béatrice Ducastel1, Laurent Cazes2, Salvador Jordana3, Martí Bayer-Raich3, Trevor Osorno4, John Frederick Devlin4

1TOTAL S.A. - PERL, 64140, Lacq – France

2TOTAL S.A. – CSTJF, 64000, Pau – France

3Amphos 21 Consulting S.L. – 08019, Barcelona – Spain

4University of Kansas, Lawrence, KS66045 – USA

*Contact author: [email protected], +33559673846

Keywords: groundwater dynamic, mass flux, innovative tools, point velocity probe, in-situ monitoring

Abstract

In the practice of environmental monitoring of industrial sites, groundwater flow parameters, such as direction and magnitude of water velocity, are typically deduced from indirect measurements made at several piezometers. Groundwater quality is generally assessed by sampling the whole column of water inside each screened well to provide one concentration value per piezometer location. These flow and concentration values give a global “static” image of contaminant distributions and trajectories with huge spatial and temporal uncertainties that can neither characterize mass discharges dynamically nor accurately. TOTAL R&D Subsurface Environmental team is challenging this classical approach with an innovative way of dynamic characterization of groundwater in shallow aquifer. The current study aims at optimizing the tools and methodologies for: first, directly measuring groundwater velocities in multilevel fashion in each piezometer and second, determining the depth-specific and total potential fluxes of dissolved contaminants in a shallow aquifer. Lab-scale experiments were conducted in a 1.5 m length sandbox, equipped with a 80 mm diameter screened well and filled with well-characterized sand, where a range of controlled flowrates were applied. Numerical modeling using COMSOL Multiphysics was performed to calibrate the sandbox and simulate some laboratory experiments. The modeling study coupled the Darcy equation in the porous media (aquifer and filter pack) and the Navier-Stockes equation within the borehole accounting for actual slot geometry (number of slots, separation and aperture). For groundwater flowrate measurements, four tools (three R&D and one commercial) were deployed and tested at different locations of the sandbox. Point Velocity Probes (PVP) were installed in direct contact with the porous media. The Colloidal borescope (CB), In-Well Point Velocity Probe (IWPVP) and Direct Velocity Tool (DVT) were deployed inside the screened borehole. In this unconsolidated porous media, PVPs presented the most reliable results with respect to direction and magnitude of groundwater flow. Experiments and modeling also showed that the structure and location of measuring tools installed in the monitoring well increased uncertainties of velocity measurement. Two additional commercial tools: the Passive Flux Meter (PFM) and i-FLUX were installed in the well to assess the controlled mass flux of organic (TAME) and inorganic (Zn)dissolved species at two different flowrates and two TAME and Zn concentrations that could be found inindustrial sites. Following laboratory testing, field-scale experiments and modeling were performed at adedicated pilot site at the French historical Lacq industrial area. Multilevel monitoring tools validated at thelab-scale were deployed at the field scale to assess groundwater flow and solute transport in highpermeable shallow unconfined aquifer. The results of these characterization tools for groundwater flow andmass flux gave encouraging results for further operating deployment. Indeed, the long-term R&D strategyis targeting in-situ, real-time, remote and cost-effective monitoring solutions for operating industrial sitesand facing challenges of heterogeneous geologies, transition areas in streambed, and estuarial andshoreline environments.

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Introduction

During environmental monitoring of industrial sites, groundwater flow parameters, such as direction and magnitude of water velocity, are classically deduced from measurements of water head gradient between several piezometers. Tracer tests, pumping tests or dilution tests can provide additional information on the dynamics of the targeted aquifer, indirectly. Then, groundwater is generally sampled across the whole water column of each screened well to provide one concentration value per piezometer location. These flow and concentration values give a global “static” image of contaminant distributions and trajectories in the aquifer, with huge uncertainties in time and space scales and therefore in any estimate of dynamic mass discharge. TOTAL R&D Subsurface Environmental team is challenging this classical approach with an innovative way of dynamic characterization of groundwater in shallow aquifers. Mass flux and mass discharge have been well studied and documented by authors for 20 years (Einarson and Mackay, 2001; Expert Panel on DNAPL Remediation, 2003; Bayer-Raich et al., 2004; Annable et al., 2005, Devlin et al., 2009, Verreydt et al., 2017; Michel et al., 2017). In 2010, the Interstate Technology & Regulatory Council (ITRC) published a dedicated technical report on mass flux and mass discharge assessment for a better understanding of groundwater and pollutant plume dynamics (ITRC, 2010). The authors noted that mass discharge and flux estimates quantify source or plume strength at a given time and location. Consideration of the strength of a source or solute plume (i.e., the contaminant mass moving in the groundwater per unit of time) improves evaluation of natural attenuation and assessment of risks posed by contamination to downgradient receptors, such as supply wells or surface water bodies. In addition to defining a source strength and plume attenuation rate, mass flux estimates – particularly if determined using transects – can identify areas of a cross-sectional plane through which the majority of the contaminant mass is moving. This information is valuable in virtually all aspects of contaminated site management. ITRC also defined mass flux as the rate measurement specific to a defined area, which is usually a subset of a plume cross section and expressed as mass/time/area (e.g., g/d/m2). The overall mass discharge crossing a boundary of concern can then be determined as an integrated mass flux estimate (i.e., the sum of all mass flux measures across an entire plume). Mass discharge is therefore expressed as mass/time (e.g., g/d) (ITRC, 2010, figure 1). The current studies are mainly focused on surface source of pollutants linked to shallow industrial activities. This approach could also be applied to environmental subsurface management of oil and gas activities or CO2 geological storage (Credoz et al., 2018). In the field, this approach will help better assess the migration of pollutants with consideration of the aquifer heterogeneities that can sometimes create variations in groundwater velocities of an order of magnitude or more over very short distances. It can also identify relevant stratigraphic differences of grain size distribution and mineral composition of the layers that may affect the mobility of contaminant plume (i.e. presence of clay lens, organic matter, etc.) (Allen-King et al., 2015). Finally, this optimization of the conceptual model and the risk assessment will benefit remediation plans and treatment system designs, reducing the global cost of the treatment.

Figure 1: On the left, example of mass flux assessment along an heterogeneous stratigraphic column: for the same concentration (C) and groundwater gradient (i) into the three layers, mass fluxes vary and therefore, the risks differ significantly because of variations in the hydraulic conductivity. On the right, Spatial variation in mass flux across a contaminant plume (ITRC, 2010 and courtesy of HydroGeoLogic Inc.).

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Currently, the literature describes several different characterization techniques for estimating mass flux and mass discharge (ITRC, 2010). Mass fluxes can be obtained by an integrative measurement of groundwater flow and integrative sampling and analysis of pollutants concentrations with passive cartridges deployed into monitoring wells for a dedicated duration (days to weeks) (Annable et al., 2005; ITRC, 2010; Verreydt et al., 2017, Michel et al., 2017). Alternatively, the monitoring of pumped effluents during Integral Pumping Tests may be used to estimate mass discharges within the total volume of aquifer sampled (Teutsch et al., 2000; Bayer-Raich et al., 2004). Another approach is the direct measurement of velocity in porous and unconsolidated media from dedicated boreholes allowed to collapse against the measurement devices (Devlin et al., 2009; Gibson and Devlin, 2018) or into non-dedicated monitoring wells (Kearl, 1997; Essouayed et al., 2018; Osorno et al., 2018). These data are later combined with groundwater sampling and classical analysis of pollutants concentrations in the lab to obtain mass discharge estimates.

For 3 years, TOTAL R&D is supporting the internal ORCAD project (Online and Real-time Characterization of Aquifer Dynamic) which focuses on the dynamic and multilevel characterization of the subsurface environmental onshore compartment. A series of technologies are currently in development and being examined experimentally in TOTAL R&D and partners’ facilities (academics, research centers, start-ups and industries). These tools and methods aim at identifying groundwater dynamics in multilevel fashion, i.e., flow velocity as a function of time and depth. A focus is also made on in situ or onsite analysis ofpertinent organic and inorganic geochemical parameters to, finally, assess mass flux and mass discharge.The main effort is put on the identification of discriminating parameters that, by themselves or together,help distinguish potential irregularities from natural variations.

Materials and methods

Experiments have been performed at Pôle d’Etudes et de Recherches de Lacq (PERL), a TOTAL Research Center located in the South West of France, at the lab scale in a metric sandbox tank equipped with one screened well, and at a multimetric industrial pilot scale site equipped with several monitoring wells. These two experimental scales are important to assess and upgrade the tools provided commercially, and further the development of non-commercial tools in order to reach enough maturity (Technology Readiness Level - TRL > 5) before deploying them at industrial and operated sites. The sandbox tank (Figure 2) has beencarefully filled with homogeneous and well-characterised sand to simulate horizontal groundwater flow atthe metric scale. The sandbox tank dimensions are: 150 cm (L) x 50 cm (l) x 50 cm (h) with two 10-cm largeupstream and downstream tanks. The total volume of pore water is 150 liters (L) with an average porosityof 0.4 for the sandy matrix. A monitoring borehole has been installed 1-m from the upstream tank with a80-mm internal diameter. CTD probes have been buried at different locations of the sandbox to measureconductivity, temperature and depth during the groundwater flow experiments. Flow rate is controlled bytwo pumps located upstream and downstream and connected to a buffering tank. The buffering tank canbe filled with an aqueous solution that could contain either, NaCl, inorganic (Zn) or organic (TAME) aqueousspecies for the dedicated tests. Daily to weekly experiments have been conducted in this facility. Numericalmodeling of the groundwater flow in the porous media and into the monitoring borehole of the sandbox hasalso been performed using the COMSOL Multiphysics capabilities (COMSOL, 2012) to obtain the velocityfield within the screened borehole (Bayer-Raich et al., 2018). It was necessary to solve the coupled problemby solving Stokes equation in the open spaces (open borehole) as in Sano (1983) for the homogeneouscase (without rings) and Brinkman-Darcy equation in the porous media surrounding the borehole. Using 3Dnumerical simulations, the influence the geometry of the well screen and the presence of in-hole measuringdevices may be analyzed.

After validation of the tools and methodology at the sandbox scale, field tests were performed at the pilot scale site on the Lacq industrial property. PERL owns a 5.5-ha pilot area named Pilots Platform of Lacq to perform R&D actions at an industrial scale. TOTAL’s ADYCHATS pilot site (for Aquifer Dynamic CHAracterization Tools System) contains 20 monitoring wells of 3 to 4 inches of diameter (Figure 3) and 10 Point Velocity Probes (PVP). PVPs were installed at five locations and two depths in the alluvium where heterogeneous sediments were expected (from silty and clayey level to gravel layer deeper). ADYCHATS projects at the pilot site aim at performing R&D in the shallow groundwater monitoring zone for subsurface

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environmental management of conventional O&G and carbon capture, utilization and storage (CCUS) activities.

Figure 2: Sandbox tank and other dedicated devices for lab-scale testing groundwater monitoring tools at the PERL Research Center.

Figure 3 : Monitoring wells on the ADYCHATS pilot site in Lacq and mapping of average seepage velocity vector measured by PVP (in porous media) and IWPVP (into the screened well) at two depths. The blue narrow stands for the deep measurement and red narrow for the shallow one.

Among the six tools tested at PERL (Table 1), four tools were deployed in the sandbox to measure groundwater flow velocity: one in direct contact with the porous medium and three in monitoring wells.

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Table 1: Groundwater monitoring tools deployed at PERL labs (Sandbox) and industrial pilot site.

Tools Reference Measurement Location of deployment

PERL Lab-Sandbox

ADYCHATS pilot site

Point Velocity Probe (PVP)

Gibson et Devlin, 2018

Seepage velocity Porous and unconsolidated media

Done: 4 velocities

Done:10 locations

Direct Velocity Tool (DVT)

Essouayed et al. 2018

Darcy flux Monitoring well Done: 2 velocities

In progress

In-Well Point Velocity Probe (IWPVP)

Osorno et al. 2018

Darcy flux Monitoring well Done: 7 velocities

Done in 1 well

Colloidal borescope (CB)

Kearl, 1997 Darcy flux Monitoring well Done: 5 velocities

Done

Passive Flux Meter (PFM)

Annable et al. 2005

Mass flux: average Darcy flux and concentration

Monitoring well Done: 2 velocities

In progress

i-FLUX Verreydt et al. 2017

Mass flux: average Darcy flux and concentration

Monitoring well Done: 2 velocities

In progress

The Point Velocity Probe (PVP, Figure 4-a) aims at measuring groundwater flow velocity (direction, and magnitude) in shallow aquifers based on mini tracer tests and tracer displacement detection (Devlin et al., 2009; Gibson and Devlin, 2018). The probe is installed directly in contact with the aquifer material. Tracer injection, displacement and detection are then controlled from the surface by the operator. Some recent advancements performed at PERL in collaboration with the University of Kansas provide for groundwater sampling and surface geochemical analysis at the location of the velocity measurement with PVP. The In-Well PVP (IWPVP; Osorno et al., 2018, Figure 4-c) is deployed in a monitoring well at the desired depth. Horizontal flow through the probe is promoted by a series of brushes, which serve as aids to control flow in and around the probe. Displacement of an injected tracer pulse is detected, using the same technology as the PVP, and is determined by the groundwater flow in the screened well. The Direct Velocity Tool (DVT; Essouayed et al., 2018, Figure 4-b) is based on the same physics of saline tracer displacement detection but the tracer trajectory is delimitated by a dedicated circuit in the tool. The colloidal borescope (CB; Kearl, 1997, Figure 4-d) measures groundwater velocity based on acquisition of several colloidal lateral displacement across a dedicated window of recording. Only the PVP, IWPVP and CB have been deployed at ADYCHAT pilot scale. For integrative mass flux measurements, two more innovative tools, the Passive Flux Meter (PFM; Annable et al., 2005, Figure 4-e) and the iFLUX (Verreydt et al., 2017, Figure 4-f) have been deployed in the sandbox, only. These two devices are composed of cartridges containing sorbents (charcoal) for velocity measurement by alcohol elution and for sorption of organic dissolved compounds for PFM and iFLUX and inorganic dissolved compounds only for iFLUX.

Results

Experiments in the sandbox were designed to resemble groundwater flow velocities found on the ADYCHATS pilot site, with average Darcy fluxes of approximately 0.5 to 2 m.day-1. Dedicated experiments were performed for PVP measurements in the sandbox for four different Darcy fluxes controlled by the operator: 16, 25, 213 and 232 cm.day-1. The average Darcy fluxes (assuming a porosity of 0.40) and standard deviation measured by PVP were 23 ± 2.3, 30 ± 3, 287 ± 35 and 354 ± 95 cm.day-1 respectively. Direction and sense of the flow were also defined for three different rotation angles of the probe (angle between main flow direction and first injection port) during each experiment and the resulting relative errors were very satisfactory: 34 ± 10°, 90 ± 0° and 41 ± 4 ° respectively.

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a)

b)

c)

d)

e) f)

Figure 4: The six tools tested: a) Point Velocity Probe (PVP), b) Direct Velocity Tool (DVT from Essouayed et al., 2018), c) In-Well Point Velocity Probe (IWPVP), d) Colloidal Borescope (CB), e) Passive Flux Meter (PFM), f) iFLUX.

Simultaneous measurements were performed with a PVP in the porous media and the DVT in the monitoring well of the sandbox (Essouayed et al., 2018). For two imposed Darcy fluxes of 16.0 and 26.0 cm.day-1, measured values by PVP were 23 ± 2.3 and 30 ± 3 cm.day-1 (Darcy flux obtained from seepage velocity with an average assumed porosity of 0.40). The average measured Darcy fluxes and standard deviations by DVT were 25.6 ± 1 and 31 ± 6 cm.day-1 respectively. The IWPVP was deployed up to 7 times in the screened well of the sandbox with up to 5 replicates for each experiment with dedicated controlled flowrates. Average seepage velocities obtained with IWPVP and standard deviations were from 107.35 ± 3.70 cm.day-1 up to 3544.98 ± 96.47 cm.day-1 for controlled seepage velocities from 35.79 up to 787.40 cm.day-1. This was done to define a calibration line and to obtain calibration factor of 4.6 used to correct subsequent assays at the PERL labs or at the ADYCHATS pilot site. This factor illustrates the offset between the seepage value obtained with IWPVP in the probe body (located inside the screened well) compared to the controlled seepage velocity imposed by the operator of the sandbox (porosity of 0.4).

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The colloidal borescope was deployed several times with five different flowrates and replicates in the sandbox. It did not produce stable signals of velocity and direction over the times of the tests, so no coherent values of Darcy flux could be obtained to compared to the imposed experimental conditions.

Five experiments were performed with two PFM cartridges provided by Enviroflux, one for Darcy flux measurement (based on alcohol elution), another for organic compounds (based on sorption of tert-amyl methyl ether, TAME). The two cartridges were deployed into the screened well during a 48h measurement period with two imposed Darcy fluxes: 115.2 and 244.8 cm.day-1 and two imposed TAME concentrations: 5 and 10 mg.L-1. Average measured PFM Darcy fluxes were one order of magnitude lower than imposed ones: 15.0 ± 0.5 and 15.3 ± 1.6 cm.day-1. Average measured TAME concentrations fitted well with the imposed ones: 4.8 ± 0.3 and 7.3 ± 4.3 mg.L-1. Overall, PFM estimates of TAME Mass flux were more than one order of magnitude lower than actual TAME Mass fluxes in the sandbox.

Eight experiments were performed with three types of cartridges provided by iFLUX: one for Darcy flux measurement (based on alcohol elution), one for organic compounds (based on sorption of tert-amyl methyl ether, TAME) and one for inorganic compounds (based on zinc sorption). Up to two cartridges were deployed simultaneously into the screened well of the sandbox during a 72 h measurement period with two imposed Darcy fluxes: 52.0 and 144.0 cm.day-1 as well as two imposed TAME or Zn concentrations : 1 and 10 mg.L-1. Average measured iFLUX Darcy fluxes underestimated the imposed ones by a factor of 2 at both pumping rates: 36.0 ± 0 and 64.0 ± 4.2 cm.day-1. Average measured TAME concentrations fitted with the imposed ones: 1.5 ± 0.7 and 6.5 ± 6.4 mg.L-1 respectively. Average measured Zn concentrations fitted with the imposed ones: 1.0 ± 0 and 22.8 ± 19.0 mg.L-1 respectively.

At the ADYCHATS pilot site, the PVP and IWPVP were deployed successfully and results are shown on the site map (figure 3-b). PVP results indicated the presence of two hydrostratigraphic units. Shallow PVPs indicated a dominant flow direction to the west-southwest at 1 to 2 m.day-1. Deep PVPs indicate a dominant flow direction to the south-southwest at 1.3 to 3 m.day-1. All PVPs measured a significant component of vertical flow, downward. IWPVP results generally duplicated these findings, showing a dominant flow direction to the south-southwest below 4.8 m bgs at 2.4 to 3.7 m.day-1 and a dominant flow direction to the southwest-west above 4.8 m bgs at 0.3 to 0.9 m.day-1. The colloidal borescope did not give stable signals in any of several campaigns that could lead to satisfactory estimates of Darcy fluxes.

Discussion

The six tools presented in this study give the magnitude of the groundwater flow. As pointed out many times (e.g. Haitjema & Anderson, 2016; Bayer-Raich et al., 2018), care must be taken to avoid confusion between Darcy flux and seepage velocity measurements or calculations. The PVP is the unique tool that can give a seepage velocity, without calibration, since it is deployed in direct contact with the porous media, whereas other velocity measurement devices are generally deployed inside screened wells. An average porosity of the sandbox was used to define the Darcy fluxes obtained by measurement with PVP. At the same time, an calibration factor (arbitrarily defined as 2) was applied for the calculation of Darcy fluxes obtained by DVT.

Assessment of the horizontal direction of flow (x,y) is a key point to better assess the local flow and transport in heterogeneous media of potential pollutants. As stated by their authors, four of the six tools tested in this study are able to provide the horizontal direction (x,y) at the same time as the magnitude of the velocity. This statement has only been validated for PVP and IWPVP during the pilot site survey. Additionally, the PVP can provide information on the 3-D seepage velocity vectors (x,y,z), as it did at ADYCHATS pilot site, and as previously documented by Schillig et al. (2014). The 3-D nature of PVP measurements allows the user to obtain the potential vertical flow between two distinct layers of the studied aquifer.

Experiments in the sandbox were performed for Darcy fluxes of 0.5 to 2 m.day-1, which approximated those found in the aquifer of the Lacq Basin. Some tools tested may be out of their velocity range of deployment

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as stated by their respective authors. PVPs installed using the direct-push tool in unconsolidated geology are expected to better approximate the instantaneous seepage velocity and avoid additional uncertainties caused by slotted well screen, gravel pack and resulting alpha factor (Sano, 1983; Bayer-Raich et al., 2018). However passive tools like PFM and iFLUX will give a complementary information for the long-term management of pollution as average values of flow and concentrations are obtained on dedicated period of deployment (several weeks) instead of instantaneous measurements.

Conclusion

This study supports a dynamic approach to aquifer characterisation, instead of the classical static approach, to better characterize an aquifer impacted by industrial activities. Six tools were tested in a controlled sandbox to obtain flow and mass fluxes of organic and inorganic dissolved species in groundwater. Three tools (PVP, DVT, IWPVP) showed promise in terms of instantaneous velocity measurement (minutes to hours) by obtaining the direct seepage velocity (PVP) or the average Darcy flux (DVT, IWPVP). The PVP showed the greatest promise as it offers a 3-D assessment of groundwater flow and less uncertainties with the limitation that it can only be deployed in an unconsolidated geology (direct-push), currently. Contrary to the PVP, the DVT and IWPVP are very suitable for deployment in existing monitoring boreholes, subject to the known uncertainties induced by the slots of the screened well and the influence of a gravel pack. Two additional integrative and passive tools were deployed for a short time period to obtain mass fluxes of organic (PFM) and organic/inorganic (iFLUX) solutes. They offer a complementary assessment of the flow field and additional mass flux assessment for a longer term (weeks). Some adjustments have to be made during the deployment to better constrain the velocity and concentration measurement at the lab scale with the known limitations of a metric sandbox and its boundary conditions. Field deployment of these tools will help at better understanding the advantages and uncertainties of these devices in the future. Repeated campaigns of measurements in different geology and geochemistry would help to better define the optimum range of operational use, taking into account geophysical characterization which would optimize the stratigraphic description of the studied media. Sites with high groundwater velocities (Darcy fluxes of 0.5 to 2 m.day-1) such as the Lacq Basin site, complicates the measurement accuracy of mass flux which remains a challenge that needs to be addressed. Multiphysics modeling remains an important tool to predict the effects of a studied tool on velocity field inside the porous media or in the screened well and recalculate the induced calibration factors. Based on the preliminary results of this work, some of these tools must be considered for a better definition of the conceptual model of a studied site and a better assessment of contaminant transport dynamics. These instruments and techniques show promise to reveal the evolution of surface and subsurface activities of production, remediation, natural attenuation and restoration throughout the life of an industrial site.

References

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POLLUTION FLUX MEASUREMENTS AS PART OF RISK ASSESSMENT FOR THE SPREADING OF GROUNDWATER POLLUTION: EXPERIENCES FROM PILOT PROJECTS

Bert Van Goidsenhoven1, Kurt Bouckenooghe², Karel Biesemans3, Isabelle T’Seyen4

1Public Flemish Waste Agency (OVAM), Stationsstraat 110, Mechelen, Belgium

2Antea Group, Roderveldlaan 1, Antwerpen, Belgium

3Arcadis, Koningsstraat 80, Brussel, Belgium

4Witteveen+Bos, Posthoflei 5-1, Antwerpen, Belgium

Corresponding author email: [email protected]

Keywords: flux measurements, passive sampling

Introduction – risk assessment and flux measurements

Carrying out a thorough characterization of a contamination together with a well-founded risk assessment forms the basis of risk-based management of groundwater contamination.

Soil contamination poses a risk of spreading when receptors are negatively affected or when there is a risk that receptors may be adversely affected in the future. Receptors of groundwater pollution can be wells, surface water, water extraction areas or groundwater dependent ecosystems.

The characterization of the pollutant flux that is released from the total waste load and that threatens a receptor is a very relevant criterion in the context of risk assessment, remediation evaluation and follow-up. In the standard procedures for descriptive soil research as they exist today in Flanders, no explicit attention is paid to the pollutant flux.

An evaluation based on only pollutant concentrations that were determined with active sampling has the important shortcoming that the resulting pollutant fluxes often do not correspond with the actual pollutant fluxes. Concentrations measured with active samples are after all a snapshot of one specific moment in time. For certain pollutants such as chlorinated solvents, samples sometimes show strong fluctuations in measured concentrations. Passive sampling techniques for groundwater allow samples to be taken over a long period of time. In this way a time-averaged picture of the concentrations can be obtained. In combination with a direct or indirect determination of the groundwater flux, the pollutant flux can be derived.

Pilot projects with flux measurements

After a standard procedure for passive sampling techniques for groundwater was established, OVAM launched a call for soil experts to present pilot projects on the use of flux measurements. The goal of these pilot projects was to collect knowledge and experiences on the use pollutant fluxes in soil investigation. An open call was launched and three proposals were withheld, submitted by respectively Antea, Arcadis and Witteveen+Bos. By mere coincidence (and somehow unfortunately), the three experts proposed the same technique for flux measurements i.e. the iFlux samplers.

The three pilot projects treated diverse types of pollution in different geological conditions. The results of flux measurements were compared with results of conventional sampling, keeping in mind that these different techniques measure different aspects. This paper does not describe the complete pilot projects but highlights some of the aspects that were observed.

The first pilot project was carried out on the site of a former sedimentation basin of a plant for surface treatment of metals. The basin has been removed since 30 years but a pollution with chlorinated solvents (VOC) and nickel remains. The geology consist of coarse to fine sands and groundwater is at 8 m between beneath surface level (m-sl). In order to assess whether the groundwater pollution is threatening nearby private groundwater extraction wells flux measurements are used in combination

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with a groundwater model. Two control planes with clusters of wells on 3 different depths were installed in the flow path of the groundwater. The results of the control plane closest to the source zone are given in table 1.

Table 1. Pollutant fluxes in control plane 1 (DL = detection limit, W = water flux cartridge, O = volatile organic components cartridge, M = heavy metal cartridge)

The results show that nickel is in general concentrated in the upper layer, the flux diminishing from 48 mg/m²/day to 2 to 5 mg/m²/day with depth. The water cartridge in 45A shows a divergent result of 2 cm/day compared to 9 to 22 cm/day in the other cartridges. The reason of this deviation is not clear. There is no correlation of the results of heavy

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