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Contract n° SSPI-2004-006538

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Research for Policy Support

D10: Impact of hydrogeological conditions on pollutant behaviour in groundwater and related ecosystems.

Volume 1

Due date of deliverable: March 2006

Actual submission date: May 2006 Start date of the project: 1st January 2005 Duration: 2 years Organisation name of lead contractor for this deliverable: BRGM

Project co-funded by the European Commission within the Sixth Framework Programme (2002-2006)

Dissimination level PU Public PP Restricted to other programme participants (including the Commission Services) RE Restricted to a group specified by the consortium (including the Commission

Services)

CO Confidential, only for members of the consortium (including the Commission Services)

D10 : Hydrogeological conditions - volume 1

Deliverable D10 3 / 158

Preface

This report is a deliverable (deliverable D10) of the BRIDGE project, a specific targeted research project (STREP) of the 6th EU RTD Framework Programme belonging to the Scientific Support Policies Priority. In particular, it is the final deliverable of Workpackage 2 (WP2) “Study of groundwater characteristics”. Dealing with the impact of hydro-geological conditions on pollutant behaviour in groundwater and related ecosystems, this report aims to synthesise the current state of the art of the knowledge throughout Europe. Expert hydrogeologists from different European countries and associated countries contributed to this synthesis through preparation of some sections, replies to questionnaires or active participation in a workshop that took place in Orléans (France). This report has a twofold objective: In a first step it provides part of the information necessary to develop recommendations on a methodology and criteria for an European approach on how to establish environmental thresholds for groundwater bodies. In a later step the same information will be useful for the application of methodology.



General Content Volume 1 Chapter 1: Introduction Chapter 2: Techniques used to collect and process the data Chapter 3: Methodologies used for GWB delineation

Introduction General comment Delineation of groundwater bodies Aquifer Typology Interaction with surface waters Assessment of pressures Assessment of vulnerability Risk assessment Open questions Concluding Remarks References of chapter 3

Chapter 4: Concepts for characterisation of aquifer regarding transport and attenuation

of pollutants

Generalities Typology of aquifers Characterisation of attenuation between source of pollution and receptor

Chapter 5: Hydrogeological processes

Importance of aquifers in the context of the WFD Concept of the aquifer control on pollutants Attenuation of the pollutants according to aquifer typology Synthesis and perspective of use of information on hydrogeological processes Appendices: in a separate volume

Chapter 6: Natural Background Levels. State of the art and review of existing

methodologies Introduction Methodologies to determine the natural background level State of the art on natural background levels Conclusions References Appendices: in a separate volume


6 / 158 Deliverable D10

Volume 2 Chapter 7: Groundwater/surface water interactions

Introduction Aim Types of surface water systems Chemical/substance concerns Processes and controls at catchment scale Processes and controls at local scale Discussion Conclusions References

Chapter 8: Groundwater/dependant terrestrial ecosystems interactions

Introduction Aim Types of water dependent terrestrial ecosystems Linking landscape location and water transfer mechanisms Processes and controls on the groundwater system and the GWDTE. Chemical /substance concerns Review of methods Discussion Conclusions

Volume 3 Chapter 9: Impact of quantitative alteration on groundwater quality

What means change of quantitative status? Synthesis Quantity related aquifer responses and triggered processes with impact on groundwater quality Matrix Actions/quantity impact/triggered processes/quality parameter influenced Assessment of quantitative impacts on quality Annex Task 2.3 partnerships: contribution of each partner



Content of Volume 1

Chapter 1 - Introduction.......................................................................................................11

Chapter 2 - Techniques used to collect and process the data ........................................15

Chapter 3 - Methodologies used for GWB delineation .....................................................19 3.1. Introduction .........................................................................................................21 3.2. General comment ...............................................................................................21 3.3. Delineation of groundwater bodies .....................................................................22 3.4. Aquifer classification ...........................................................................................22 3.5. Interaction with surface waters ...........................................................................26

3.5.1. General remarks ......................................................................................26 3.5.2. Specific comments on responses ............................................................26

3.6. Assessment of pressures....................................................................................27 3.7. Assessment of vulnerability ................................................................................27 3.8. Risk assessment.................................................................................................30

3.8.1. General remarks ......................................................................................30 3.8.2. Methodology ............................................................................................30 3.8.3. Preliminary evaluation of aquifers being at risk .......................................34

3.9. Open questions...................................................................................................34 3.10. Concluding remarks ..........................................................................................34

3.10.1. Definition of the terms............................................................................34 3.10.2. Current state of RBDs and GWBs .........................................................35 3.10.3. Typology of GWBs.................................................................................35 3.10.4. Numbers and dimensions of RBDs........................................................35 3.10.5. Percentage of surface of the country covered by the GWB...................35 3.10.6. Numbers and dimensions of the GWB ..................................................35 3.10.7. Assessment of risk not to reach good status of environmental

objectives in 2015 ..................................................................................36 3.11. References........................................................................................................38

Chapter 4 - Concepts for characterisation of aquifer regarding transport and fate of pollutants .....................................................................................................47

4.1. Generalities.........................................................................................................49 4.2. Typology for hydrogeochemical characterisation of aquifers..............................49 4.3. Characterisation of attenuation between source of pollution and receptor .........51



Chapter 5 - Hydrogeological processes.............................................................................55 5.1. Importance of aquifers in the context of the WFD...............................................58 5.2. Concept of the aquifer control on pollutants .......................................................59

5.2.1. General characteristics influencing both the flow system and physico-chemical processes ..................................................................60

5.2.2. Residence time ........................................................................................61 5.2.3. Flow processes – attenuation by dilution.................................................66 5.2.4. Physico-bio-chemical attenuation in aquifers ..........................................68 5.2.5. Hydrogeochemical background ...............................................................70

5.3. Aquifer typology and attenuation of pollutants ....................................................70 5.3.1. Principle and objectives of the typology...................................................70 5.3.2. Sandstones - siltstones aquifers..............................................................71 5.3.3. Sands and gravels aquifers .....................................................................73 5.3.4. Limestones (karstic/non karstic) aquifers ................................................76 5.3.5. Chalk aquifers..........................................................................................78 5.3.6. Schists and Shales aquifers ....................................................................79 5.3.7. Crystalline aquifers ..................................................................................80 5.3.8. Volcanic rocks..........................................................................................81

5.4. References .........................................................................................................82

Chapter 6 - Natural background levels. State of the art and review of existing methodologies .................................................................................................85

6.1. Introduction .........................................................................................................87 6.2. Approaches to determine Natural Background Levels........................................88

6.2.1. Main approaches ...................................................................................88 6.2.2. National approaches (FR, D, BL, LT).....................................................89 6.2.3. Local scale approaches (UK, DK, BL-Flanders, EE, NL, HU, PT) .........95 6.2.4. The Baseline project ..............................................................................97 6.2.5. Conclusion .............................................................................................99

6.3. Typology of aquifers ...........................................................................................99 6.3.1. Description of the typology.....................................................................99 6.3.2. Approach to define the “typical natural chemical composition” of

each type of aquifer .............................................................................100 6.4. Natural Background Levels in Limestones aquifers..........................................102

6.4.1. Major elements ....................................................................................102 6.4.2. Trace elements ....................................................................................105

6.5. Natural background levels in chalk aquifers .....................................................111 6.5.1. Major elements ....................................................................................111 6.5.2. Trace elements ....................................................................................111

6.6. Natural background levels in sands and gravels aquifers.................................113




6.7. Natural background levels of sandstones aquifers ...........................................120 6.7.1. Major elements ....................................................................................120 6.7.2. Trace elements ....................................................................................122

6.8. Natural background levels in clayey and marly aquifers...................................125 6.9. Natural background levels in crystalline basement rocks aquifers ...................125


6.10. Natural background levels in schists and shales aquifers...............................130 6.11. Natural background levels in volcanic aquifers ...............................................130

6.11.1. Trace elements ....................................................................................132 6.12. Synthesis.........................................................................................................132


6.13. Saline influence...............................................................................................134 6.13.1 Salt water..............................................................................................134 6.13.2. Mixing of salt water with fresh water....................................................141 6.13.3. Tools for detecting sources of salinization...........................................142

6.14. References......................................................................................................153



Chapter 1

Introduction



The future European Groundwater Directive, a draft proposal of which has been adopted by the Commission in its final form on January 23 20061 is a daughter directive of the Water Framework Directive (2000/60/EC). It represents a comprehensive piece of legislation that will set out clear quality objectives for all groundwaters in Europe. Criteria for the assessment of the chemical status of groundwater are partly based on existing Community quality standards (nitrates, pesticides and biocides) but Member States are required to identify pollutants and threshold values that are representative of groundwater bodies found as being at risk, in accordance with the analysis of pressures and impacts carried out under the WFD. In the light of the above, the BRIDGE project, EC 6th Framework Programme project, belonging to the Scientific Support policies priority, has for main objective to derive a plausible general approach, how to structure relevant criteria appropriately with the aim to set representative groundwater threshold values at national river basin district or groundwater body level in a scientifically sound way . This report is a deliverable of the BRIDGE project, prepared to gather part of scientific outputs which could be used to set out thresholds values for pollutants within groundwater. Its aim consists in sharing knowledge on groundwater body/aquifer characteristics of relevance regarding pollutants behaviour and discussed within the frame of Workpackage 2 “Study of groundwater characteristics“ of the project. It is the hydrogeological counterpart of Workpage 1 focusing on geochemistry and pollutants properties. The present document summarises key elements of groundwater body/aquifer characterisation relevant to pollutants behaviour as described and discussed by partners from 16 European and associated countries (Austria, Belgium, Bulgaria, Denmark, Estonia, Finland, France, Germany, Hungary, Italy, Lithuania, the Netherlands, Poland, Portugal, Spain and United Kingdom). This report synthesizes partners contribution within the frame of WP2 and comprises 9 Chapters. After this introduction, Chapter 2 presents the approach followed to collect all data on groundwater bodies’ characteristics. Chapter 3 briefly points out the broad range of approaches adopted by Member States delineation and characterisation of groundwater bodies. Chapter 4 sets keys elements for characterizing aquifer with respect to pollutant transport and attenuation. Chapter 5 describes the properties of different European hydrogeological settings with regards to geochemical properties of pollutants. Chapter 6 addresses the methodologies developed within Member state for determination of Natural Background levels of chemical elements within groundwater with a discussion of the results obtained at abovementioned European hydrogeological settings. One key purpose of the WFD being to prevent further deterioration and to protect and enhance the status of aquatic ecosystems, and with regard to their water needs, terrestrial ecosystems and wetlands directly depending on the aquatic ecosystems, chapters 7 and 8 deal with the links between groundwater and surface water and dependent terrestrial ecosystems. Finally, Chapter 9 is devoted to the impact on groundwater quality of changes in groundwater quantitative status.

1 COM(2003)550



Chapter 2

Techniques used to collect and process the data



In view of the important number of partners (18) involved within WP2; questionnaires were largely used to collect available scientific data. Such a way enabled to collect data from a wide variety of hydrogeological settings so as provided description is representative of European aquifers. Firstly, a general questionnaire was intended to identify the current knowledge and any information gaps that must be addressed within the WP2 and record specific interests and capabilities of partners. This first questionnaire was composed of three types of questions: - Questions related to available information in countries about the processes for delineation

and initial characterisation of groundwater bodies; - Questions about the experience of partner’s group; - Questions related more explicitly to the work to be conducted by partners within WP2. This questionnaire was sent to all partners involved within WP2, who are representatives from 14 countries (Austria, Belgium, Bulgaria, Denmark, France, Estonia, Germany, Hungary, Italy, The Netherlands, Poland, Portugal, Spain and United Kingdom). Additionally, responses to questions related to available information within countries about processes for delineation and initial characterisation of groundwater bodies were delivered by 2 additional countries (Finland and Lithuania). Replies to first part of questionnaire cause writing of chapter 3. Answers to other questions serve as a basis for elaboration of content of chapters 5 to 9. Actually, this questionnaire was followed by 3 more detailed questionnaires. 1. A questionnaire on Natural Background Level (NBL) including questions about: - Methodologies developed within countries or at case study scale to determine NBL, - Main results obtained on selected aquifer by applying previously described methodologies

(Identification of major and trace elements with high concentrations, with main characteristics of concerned aquifer and hydrogeological settings.

Replies cause writing of Chapter 6 which was discussed and amended by the partners during lateral discussion (via e-mail) and during a special meeting held during WP2 workshop of September in Orléans). 13 Partners, representatives from 11 countries (Belgium, Bulgaria, Germany, Denmark, France, Estonia, Hungary, Lithuania, The Netherlands, Poland and United Kingdom) contributed to this chapter. 2. A questionnaire on contribution of groundwater to surface water and groundwater

dependent ecosystems, including questions about: - Methods or estimating flow contribution of groundwater to surface water or GW dependent

ecosystems, - Methods for estimating the attenuation of pollutants at the groundwater/surface water

interface, - Examples of application within catchments. Replies cause writing of both chapters 7 and 8, which were discussed and amended among partners via e-mail and though WP2 Workshop.



Partners involved within this activity are originally from 11 countries (Austria, Bulgaria, Germany, Denmark, Estonia, France, Hungary, Italy, Poland, Portugal, United Kingdom). 3. A questionnaire on pollutant attenuation capabilities of European aquifers, including

questions on: - Hydraulic/hydrogeologic characteristics with a potential impact on dilution of pollutants, - Characteristics of aquifers having an impact on physico/bio-geochemical attenuation (like

precipitation, adsorption, degradation…). Examples from 9 countries (Belgium, Estonia, France, Germany, Hungary, Italy, Portugal, Spain and United Kingdom) have been provided by partners. These contributions gave rise to chapter 5. However, questionnaire forms were rather late, after September meeting, and consolidated version of chapter has not been discussed and amended by partners. It is expected for D10 to be produced by March 2006. Chapter 9 dealing with impact of quantitative alteration of groundwater status on groundwater quality involves a more restricted number of partners than other activities (actually 7 countries: Belgium, Estonia, France, Spain, Poland, Portugal, and United Kingdom). Consequently, work has been structured in a different way. First, partners agreed on a list of typical situations where change of quantitative status may impact qualitative status (both human induced changes and natural/hydroclimatic changes). Secondly, each partners took in charge one or several of these typical situation and produce a description of relation between quantitative and qualitative status changes as well as the methods for detecting these relations For each of chapters 6 to 9, devoted to synthesis of scientific knowledge, the contents is partly related to partners activities/case studies but also include relevant recent literature, although the topics being much too vast to achieve an exhaustive literature review, only some significant references are cited.



Chapter 3

Methodologies used for GWB delineation

M. Normand, A. Blum, H. Pauwels, S. Urban (BRGM) Zoltán Simonffy (BME)

Ph. Meus (ULG) Carlos Martínez Navarrete, José Antonio de la Orden Gómez, Juan Grima Olmedo (IGME)

Stanislaw Witczak (DHWP/AGP) Sophie Vermooten, Jasper Griffioen (TNO) Kestutis Kadunas, Rytis Giedraitis (LGT)

Frank Wendland (FZ-Jülich), Johann-Gerhard Fritsche (HLUG)

Rüdiger Wolter (UBA-D) Andres Marandi (UT)

Alwyn Hart, Jan Hookey (EA) Klaus Hinsby (GEUS), Mette Dahl (GEUS), Kim Dahlstrom (Danish EPA)

Martin Skriver (Danish EPA), Per Rasmussen (GEUS) Andreas Scheidleder, Karin Weber, Arno Aschauer (UBA-A)

Rossitza Gorova (EEA) Gustafsson Juhani (SYKE)

Marleen Coetsiers, Kristine Walraevens (LAGH-UGent)

With the contribution from Vincent Fitzsimons (SEPA-Scotland)

Petra Snellings, Marleen Van Damme (AMINAL-Flanders)



3.1. Introduction This chapter summarises and comments responses obtained through questions related to available information about characteristics of groundwater bodies, the way they are delineated and their initial characterisation in each country, questions which were included in the first questionnaire sent to partners. This questionnaire was sent to all partners involved in WP2, representing 14 countries (Austria, Belgium, Bulgaria, Denmark, France, Estonia, Germany, Hungary, Italy, The Netherlands, Poland, Portugal, Spain and United Kingdom). Some countries have sent several responses separately according to their regional structure: Great Britain (England & Wales, Scotland) and Belgium (Flanders, Wallonia). 3.2. General comment The questionnaire content has been detailed within D8. Most questionnaires have been – partly or totally – filled out with more or less details. The responses are showing a large diversity. Of course such basic information as numbers of district and, in a broad sense, the number of groundwater bodies is reflecting the situation of each of the countries. But issues like delineation of groundwater bodies, typologies, classifications and methodologies are showing a wide range of approaches. The purpose of this report is not to describe the specific way of each country: the feedbacks are only summarising some points and the writers mostly indicate background national studies concerning particular issues. Taking into account all information contained in these documents would go beyond of the scope of this report. The questionnaire is intended to identify the current knowledge and any information gaps. Results are summarised in two sections: - Similarities and differences between the countries for some key points; - List of open points. These similarities and differences are regrouped and presented in table form. The text describing each national approach is nearly the same than the original text in the questionnaire responses in order to reduce interpretation errors. It should be noticed that the use of this kind of classification is not done in an exclusive manner. For example, Finland and England & Wales are associated as countries using a “water abstraction” based classification of groundwater bodies. This doesn’t mean that: - Other characteristics of groundwater bodies are less relevant for the approach of both

countries; - “Water abstraction” couldn’t play an important role for the delineation or the typology for

the other countries. It is only meaning that the both questionnaire responses are emphasising this point.



Otherwise, important issues like Groundwater Monitoring could not be treated because most of the responses weren’t detailed enough to try a serious comparison. 3.3. Delineation of groundwater bodies The delineation of the groundwater bodies has been treated by all countries taking into account the hydrographical limits, the hydrogeological parameters and - depending on the countries - the water use. A few particular points should however been noticed: - Some countries like France and Austria indicate that the whole territory has been assigned

to be covered by more or less shallow groundwater bodies. On the contrary in the case of Finland, the total surface of the delineated groundwater bodies covers only 4.1% of the whole territory;

- Belgium (Flanders), Estonia and Lithuania are indicating that main groundwater systems or bodies are divided in sub-units;

- Bulgaria is divided in 4 river basin directorates, but at the same time in 3 hydrogeological districts.

- Because of the geographical position of Hungary, more than half of the water bodies are partly delineated by the country borders. However they belong to only 1 River basin district.

The Netherlands indicate a classification of groundwater bodies respectively on national (type 1) and in regional (type 2) levels. 3.4. Aquifer classification The aquifer typology has been treated differently. We can broadly define two groups of countries referring to the most important classification criteria: - Water abstraction based classification (England and Wales, Finland); - Hydrogeology based classification (the others countries). The classification referring to water abstraction has been treated as follows: Countries Classification

Finland

Class I holds areas important for water supply. From these areas water is extracted and is used by water works which supply at least to 10 or more households (approximately 50 persons). Class II holds areas suitable for water supply. These aquifers are suitable for water supply, but for the time being, the areas are needed neither for the municipal water supply nor for households in the sparsely populated areas. Class III holds other groundwater areas, which need further studies to find out the suitability of the area for water supply.

United Kingdom (England and Wales)

Principal aquifers – those with significant resources which need to be managed through abstraction licensing within a Catchment Abstraction Management Strategy (CAMS) in order to prevent over-exploitation, or those with a significant role in sustaining groundwater dependent ecosystems. Secondary aquifers – which also have significant resources but with hydraulic properties which limit over-exploitation. These aquifers would not normally warrant special consideration for CAMS but may still support important abstractions and dependent ecosystems which may be subject to risks associated with pollution pressures.



Unproductive Strata – mostly limited to Tertiary and Jurassic Clays (e.g. London, Oxford, Kimmeridge Clays), which are generally unable to support abstractions greater than 10 m3/d, are unlikely to provide significant baseflow or wetland discharges, and will be considered as ‘not at risk’ without further analysis. A fourth aquifer type – significant drift aquifer has been defined and delineated from a series of workshops at regional level to show areas where significant groundwater resources occur within the drift overlying unproductive strata.

Belgium (Flanders)

Aquifer classification: an official HCOV code exists. This numerical code forms the standard hydrogeological code for the Flemish subsoil. This code is based on hydrogeology and hydrostratigraphy. A hydrogeological unit can contain layers from different lithostratigraphic units and layers that belong to the same lithostratigraphic unit can be split up over different hydrogeological units. This code is complete and unambiguous and has a chronological, hierarchical build up. The main units are: - 0000: undefined; - 0100: Quaternary Aquifer Systems; - 0200: Campine Aquifer System; - 0300: Boom Aquitard; - 0400: Oligoceen Aquifer System; - 0500: Bartonian Aquitard System; - 0600: Ledo-Paniselian-Brusselian Aquifer System; - 0700: Paniselian Aquitard System; - 0800: Ypresian Aquifer; - 0900: Ypresian Aquitard System; - 1000: Paleoceen Aquifer System; - 1100: Cretaceous Aquifer system; - 1200: Jura-Trias-Perm; - 1300: Palaeozoic Basement. Delineation of groundwater bodies: The delineation of GWB in Flanders is based on subdivision into five groundwater systems based on regional groundwater flow: - the Palaeozoic Basement System; - the Central Flemish System; - Coastal Plain and Polders System; - Bruland Chalk System; - Central Campine System; - Meuse System. These systems may be vertically superimposed. The subdivision in GWB is based on following criteria: - groundwater system boundaries; - geological boundaries / semi-permeable layers (HCOV code); - groundwater divides; - basin divides (Scheldt and Meuse basin); - salinisation boundaries; - district boundaries; - rivers; - boundary confined / unconfined part of aquifer; - isolation of problem areas (e.g. depression cone in Palaeozoic Basement

Aquifer). In Flanders an aquifer will be divided into different groundwater bodies, which is an important difference with respect to other countries where a GWB is mostly built up by different aquifers.



It should be noticed that the Class III by Finland and the so-called Unproductive Strata by England & Wales are referring to the limit of identification of groundwater bodies defined by the EU – WFD. However, this specific point doesn’t mean that the other parameters are not taken into account. Moreover, the class “Secondary aquifers” refers at least to hydrogeological characteristics (hydraulic properties). The others countries are referring more ore less to physical characters of the aquifers in appropriate manner to the national and regional situation: Countries Classification of aquifers

Austria The following distinctions were made: single GWB - groups of GWBs; shallow GWBs - deep GWBs; predominantly porous media - karst - fractured GWBs (and subtypes).

Belgium (Flanders) Hydrogeological code for the underground of Flanders (HCOV)

Belgium (Wallonia)

Although no special parameter was explicitly taken into account, it is clear that the division of the region in its different aquifers takes hydrogeological characteristics as criteria. Basic distinction was made between several typologies: fissured and karstified carbonate rocks (hercynian limestones), fissured and porous carbonate rocks (Mesozoic chalks), porous Cainozoic rocks (sands, alluvial deposits).

Bulgaria

The big variety of tectonic geological as well as topographic structures, leads to the big variety of aquifers and hydrogeological conditions prevailing in Bulgaria. With respect to collectors we can distinguish the following types – porous, karst, fissured, porous-karst, karst-fissured, porous-karst-fissured. With respect to the groundwater flow type – unconfined, semi-confined (i.e. the aquifers in the river terraces with two layered structure – lower layer – sands and gravels and pebbles and upper layer – fine sands and clayey sands and clays), confined. With respect to temperature the groundwater are cold and thermal ones. Otherwise, 7 types of aquifers are defined by their vulnerability: - First category - Highly karstified limestone rocks; - Second category - Limestone rocks without open karst forms on the surface and

alluvial-drift and drift sediments; - Third category - Alluvial deposits – river terraces , Neogene sandy deposits in

North-West Bulgaria – upper part of Neogene – the zone of recharge, Lower Cretaceous deposits – covered by loess;

- Fourth category – Neogene deposits in sandy-clayey facieses or thinly layered lake sediments – limestones;

- Fifth, sixth and seventh category – Fissured aquifers.

Denmark

An aquifer typology was also given in the Danish guidance document mentioned above. The typology is based on the most important properties of a groundwater body controlling its natural groundwater chemistry, its ability to cause different biogeochemical processes to entering pollutants, and its interactions with and thereby possible influence on surface waters and dependent terrestrial ecosystems. The following three parameters were used: - Lithology of the groundwater body, separating between siliceous and

calcareous sediments or rocks; - Contact with surface water, separating between three contact types: part of the

year (local), all year (regional) and none (deep); - Redox conditions, separating between oxidised and reduced; - This typology gives rise to 12 aquifer / groundwater body types.

Estonia Porous and fracture aquifers.

Finland There is a classification based on the geological type of the aquifer (esker, ice-marginal formation etc.) and groundwater flow pattern (anti- or synclinal).



France

The first and main level (obligatory factors) includes: - The GWB are classified, primarily according to their lithology, in the 6 following

divisions: Alluvial systems, volcanic systems, crystalline basement, dominant sedimentary rocks non-alluvial, hydraulic complex systems adapted to intensely folded mountain areas, impermeable rocks with local small aquifers, and second according to the different pressures.

- The GWB are classified according to the aquifers flow types with 3 divisions: confined, unconfined and associated unconfined and confined aquifers.

The second level of characteristics is of minor importance (optional factors, natural vulnerability). It includes: - karstification; - the presence of a border coastal area with a risk of saline intrusion; - the possibility of aggregation into only one GWB of horizontally and/or vertically

disjointed hydrogeological entities (aquifers) . These hydrogeological entities may present the same hydrogeological and pressures characteristics.

Germany

To each lithological unit have been attributed criteria as: Rock type, consolidation, conductivity, type of porosity, geochemical type of rocks. Combinations of these criteria are the base for building map titled “aquifer type” (combination between geochemical type and type of porosity) or “upper aquifer” (combination between conductivity, consolidation, geochemical type).

Hungary

Keeping geological-hydrogeological aspects in view, water bodies were designated according to the following hierarchy: - Waters in predominantly porous formations in basins:

. Cold waters: Subsurface catchment areas: - Water bodies with dominant downward flow, - Water bodies with dominant upward flow;

. Thermal waters: Water bodies by major hydrodynamic units - Karstic waters: Structural units:

. Cold waters: Water bodies by catchment areas of main groups of springs

. Thermal waters: Water bodies by major hydrodynamic units - Water in the mixed formations of mountain areas (excluding karstic waters

classified in the above group):Water bodies by structural units & surface catchment areas

Lithuania No precision about a specific typology, only inventory.

Netherlands Five types of surface geology are considered: sand, clay/peat, loess, dunes and limestones.

Poland

Poland is divided into two provinces: lowland province and mountain-upland province. In the lowland province mainly the Quaternary/Tertiary porous aquifers are dominating. In the mountain-upland province, the Mesozoic and Paleozoic systems occur. Here, the following classes are distinguished: porous-fissured, karstic-fissured, fissured, porous-karstic-fissured.

Portugal No precision about a specific typology, only inventory. Spain No precision about a specific typology, only inventory. United Kingdom (Scotland)

Typology based on following parameters: Overlying strata, transmissivity, confinement, porosity, groundwater chemistry, depth of groundwater flow, length of groundwater flow paths.

As a point to be discussed, we consider that the approaches which are using first a lithological classification and adapt it to local water use through division in sub-bodies may be the most sustainable system in the long term, avoiding major changes of delineation related to changes in water management.



3.5. Interaction with surface waters

3.5.1. General remarks One of the key points of the EU WFD is to consider at the scale of entire hydrographic catchment areas (River Basin District) the two water cycles components: surface water and groundwater bodies. These water bodies are characterised through: - Quantitative and qualitative status in the case of groundwater bodies; - Qualitative and ecologic status in the case of surface water bodies. Curiously, the WFD does not take into account the quantitative status of the surface water bodies. The link between both types of water bodies is assumed through the constraint that the exploitation of groundwater bodies mustn’t threaten the good ecological status of ecosystems (wetlands and rivers). The good environmental status concerns the qualitative but also quantitative issues. One should not forget that in low water regime (no rainfall), the rivers are exclusively recharged through the input of aquifers. Therefore, the exploitation of the aquifers will reduce this recharge component. In case of overexploitation, there is no more discharge of aquifers to rivers. This has been regularly observed in France in the department of Deux-Sèvres and Charente-Maritime. In these cases there is no adequacy between the available water resources and the production yield. In general these quantitative and qualitative interactions are often insufficiently known (few measurement or studies). Despite of the evidence of the relations, the impacts are underestimated (lack of available data). Thus in France, only 21 groundwater bodies presenting quantitative risks are listed without clearly identifying those which will present definitively a risk for the surface bodies in 2015.

3.5.2. Specific comments on responses Most commonly, ongoing studies on the interaction of surface and groundwaters focus on protected areas (Natura 2000, Habitat and Bird Directive sites, RAMSAR wetlands, national parks, etc.). Beside the quality problems, England & Wales define specific classes for rivers and lakes referring to the influence of groundwater abstraction. On the other hand Denmark introduces a detailed typology of the interaction according to the dominant water input: Precipitation dominant wetlands – Groundwater dominant wetlands – Surface water dominant wetlands. Poland and Estonia mention problems of interactions with surface water linked with mining activities. Hungary mentions the decrease of temperature of a thermal source (source Heviz) as a consequence of large water abstraction (mining activities). Otherwise, 33 biotope types (both aquatic and terrestrial ecosystem) are identified where groundwater plays a significant role on preserving their good status. 85 water bodies are concerned by these interactions.



3.6. Assessment of pressures Pressures are presented in more or less details. However a majority of countries which have provided an answer have mentioned nitrates, phosphates and pesticides as important pollution issues. Agricultural activities, but also urbanisation (leaking sewers, etc.), are pointed out as pollution sources. Typical urban and industrial pollution are also mentioned but to a lesser extent. Saline contaminations are mentioned in nearly all countries with coastal areas. At least, Bulgaria, Denmark, France, Hungary, Spain, England & Wales and Scotland are pointing out that these pressures, especially for diffuse pollution sources, have been evaluated using maps, data-banks and GIS-software. Hungary describes the following method of pressure evaluation for point source of pollution: The load of pollutants cannot be directly estimated. It is replaced by the estimation generalized hazard index, i.e. the recharge at the polluted site is multiplied by the ratio of the concentration at the receptor level (top of the aquifer used or intended to use for drinking water supply) and the threshold (standard) of the given pollutant. This simple method corresponds to recharge areas (dominantly downward flow until the top of the aquifer and downward and horizontal flow further in the aquifer). The database of the point sources provides the location, the extension of the site, type of the pollutant, information on the already polluted zone (not for all sites), and rarely the observed concentration in the groundwater. Recharge maps are available for the country. Because of the often missing starting concentration, the ratio of the real concentration and the threshold is included directly as a factor of hazard, since usually this ratio is increasing with hazard of the pollutant (e.g. for macropollutant is in the order of magnitude of 10, while for hazardous micropollutants it can reach 1000s and for others in between). 3.7. Assessment of vulnerability Utilisation of the DRASTIC software: Bulgaria, Portugal, Spain. No mention of vulnerability map for Hungaria. Otherwise, besides the general mention of vulnerability maps, several countries indicate which parameters and assessment methodology have been used:



Countries Vulnerability / maps and methodology Bulgaria In 1981 in Bulgaria was made a hard copy map on vulnerability of groundwater. The

principles for creating it are as follows: - geological and hydrogeological conditions in aquifers situated nearest to the surface

- permeability, unsaturated zone – permeability, tectonic zones, recharge conditions;

- as a basis the migration characteristics of nitrates and chlorides were used, which are not involved in adsorption processes and decay;

- the basis is the hydrogeological map of Bulgaria S 1:200,000 of 1975 and the Geological map of 1962 S 1:200,000 and the Map of Quaternary deposits and geomorphology of North Bulgaria of 1975.

7 types of aquifers concerning the vulnerability are distinguished: - First category - Highly karstified limestone rocks; - Second category - Limestone rocks without open karst forms on the surface and

alluvial-drift and drift sediments; - Third category - Alluvial deposits – river terraces , Neogene sandy deposits in

North-West Bulgaria – upper part of Neogene – the zone of recharge, Lower Cretaceous deposits – covered by loess;

- Fourth category – Neogene deposits in sandy-clayey facieses or thinly layered lake sediments – limestones;

- Fifth, Sixth and seventh category – Fissured aquifers. Denmark There has been a national analysis of vulnerability of groundwater bodies in relation to

pollution with nitrates, in which the national counties have designated nitrate-vulnerable areas based on hydrogeological parameters.

France One must clearly specify the concept of vulnerability distinguishing: - general intrinsic vulnerability of the aquifer; - specific conditions of vulnerability depending on the various types of pollutants. Attention with the application of these criteria of vulnerability to the GWB for: - various types of pollutants; - areas varying from a few tens of km² to several tens of thousands of km². Maps of vulnerability have been established since the end of the 60’s using different approaches at different scales. The current French guide (may 2003) doesn’t give explicitly a specific method to be followed. Only elements of appreciation of the aquifer vulnerability are listed. Parameters like surface slope, precipitation, impermeable covering layers, thickness of the non saturated zone, permeability of the aquifers have been evaluated. The combination of the parameters, each having a different weight factor, allows calculating a vulnerability index. The aquifer vulnerability has been included in the assessment of the risk of failing good status (comparison between chemical analysis and pressures). In that case, it concerns pesticides, nitrates, ammonium, Cl, SO4, chlorinated solvents and others pollutants if available. In 2004, the BRGM has developed a methodology based on two GIS: The existing hydrographic networks (data base CARTHAGE) and the theoretical talweg (after the Digital Model of Ground). The comparison of these two categories of information allows evaluating a broad surface run-off factor and, conversely an average infiltration factor. This result combined in particular with the parameter “depth of the non-saturated zone” gives a first approach of the intrinsic vulnerability. This method has been applied by 4 river basin agencies in the frame of on-going studies.

Germany The geological surveys and BGR (working group on WFD) have established a vulnerability map for the purposes of the WFD. More detailed vulnerability maps exist in some federal states or for special regions. The methods used are based on the evaluation of the characteristics of the covering layers.

Lithuania Main parameters included in vulnerability mapping are: depth of shallow groundwater, lithology and velocity of water movement throughout the vadose zone. In future this methodology is planned to be validated using data of surface water bodies at risk.



Poland The vulnerability map of Poland (1:500,000) should be accomplished in June 2005. (The main parameter is the travel time – Mean Residence Time MRT, of a conservative solute based on the piston flow model. MRT means exchange of total water column in the covering profile by recharge rate). Information layers of GW endangering factors (five classes) are existing in an electronic version of Detailed Hydrogeological Maps (1:50,000) covering the whole country.

Spain No vulnerability maps have been elaborated for the entire Spanish territory. But such maps exist for some specific areas. For limestone aquifers, the “COP method” is used. It is the result of multiplying three factors: O (Overlaying layers), C (flow Concentration) and P (Precipitation). Factor O refers to the capacity of the unsaturated zone to protect and filter the pollution. Factor C refers to the surface conditions that have an influence on the water flow to the maximum infiltration zones, where the possibilities for human actions are less. Factor P refers to the influence of rainfall for the transport of pollutants to the saturated zone. Factors C and P are corrective factors for the protection degree of aquifers defined in factor O.


Vulnerability is assessed by considering the following parameters: - Soils – Vulnerability maps have been produced for the whole of England and Wales.

The soil vulnerability is based on the parent rock type. The categories of vulnerability are divided in to High, Intermediate and Low for each aquifer type (Major / Minor / Non aquifers). The present maps do not take surface drift into account.

- Aquifers – The Environment Agency currently subdivides permeable strata using a classification system which is predominantly based on the ability of the strata to attenuate contaminated recharge water entering at the surface (Policy and Practice for the Protection of Groundwater, 1998). Major aquifers are the most permeable and are usually capable of supporting large abstractions. Minor aquifers are less productive but still form an important resource. Finally, formations with negligible permeability are classified as Non Aquifers.

- Source Protection Zones – Source Protection Zones are delineated around groundwater abstractions used for public water supply, food use or potable supply. Source Protection Zone I is the highest risk zone, and represents a 50-day travel time for water in the saturated zone of the aquifer to the abstraction point. Source Protection Zone II represents a 400-day travel time in the saturated zone of the aquifer to the abstraction point and Source Protection Zone III represents the total catchment area for the groundwater abstraction.

- Unsaturated Zone – The depth from the base of the proposed activity to the water table, the available depth of unsaturated zone is calculated. The unsaturated zone is considered as an important potential barrier to potential pollutants. The attenuation parameters and the depth of the unsaturated zone are taken into account.

- Post – WFD – The aquifer typology terms are changing from Major / Minor / Non – aquifers to Principal / Secondary / Significant Drift and Non-productive aquifer units. It is possible that other changes will be brought in too. The Policy and Practice for the Protection of Groundwater is due to be superseded by the Groundwater Strategy.

United Kingdom (Scotland)

Vulnerability map to describe geological pathway. Key elements are thickness & permeability of strata overlying the groundwater body. Properties of aquifer are largely mapped separately. Together they describe the vertical and horizontal pathways by which pressures can reach receptors.



Belgium (Flanders)

Groundwater vulnerability maps have been made in 1986 for each province (5 in total) in Flanders on a scale of 1:100,000. The vulnerability map can be described as a map of the degree of risk for pollution of groundwater in the upper water-bearing layer by substances that enter the soil from the ground surface. The degree of vulnerability is based on the nature of the rock and the permeability, the presence, nature and thickness of a covering layer and the thickness of the unsaturated zone. Vulnerability with respect to nitrate pollution: For the purpose of evaluating the specific vulnerability with respect to nitrate pollution, Flanders has been subdivided into hydrogeological units and zones, based on geological characteristics (e.g. thickness of Quaternary deposits, different ages of Palaeozoic clay layers). The nitrate sensitivity of each zone was assessed on the basis of the following hydrogeological characteristics: hydraulic conductivity, hydraulic gradient, degree of oxidation of the sediment during deposition, thickness of the unsaturated zone, thickness of the saturated oxic zone, absence in the sediments of effective reduction capacity (reactive organic matter or sulphides).

The table shows that most of the participants are using several parameters which allow assessing the aquifer vulnerability. Usually, this vulnerability can be represented in a synthetic map. 3.8. Risk assessment

3.8.1. General remarks The EU countries should identify the groundwater bodies which risk failing a good environmental status in 2015. In this task, they are supported by the WFD and the other EU-documents which have been especially elaborated (guidelines and methodologies). Some countries like France have elaborated their own national methodological guidelines for the risk assessment. Taking into account that the EU daughter directive is still under elaboration, the French Ministry of Environment has formulated hypotheses of work. The other EU-countries have certainly proceeded in the same way. The status of a groundwater bodies is the result of a cross-analysis of all risks (see page 35 to 37 of the guide): - Quantitative risks; - Qualitative risks. For each groundwater body the worst status case will define the general status. The methodologies are varying between the countries and certainly here an effort should be put into harmonisation.

3.8.2. Methodology Countries Methodology for Risk Assessment Bulgaria On local level Methods are available that are approved by the Minister of

Environment and waters (MoEW), as well as Methodological Instructions (1998) about the scope and content of the report on determination the damages of contamination previous to privatisation. The procedure of this evaluation concerns the liability of the government for proved previous pollution damages, i.e. the new owners “shall not be liable for ecological damage caused by past actions or lack of actions”. The process of making this report involves a two-stage procedure: - The first step is preparing a preliminary report, filling up questionnaires and forms



for identifying the contamination concerning the groundwater, the surface water, the soils and the air - by assessment on the basis of a point system (the score is calculated). Useful are the help work-schemes with list criteria for probability of the groundwater contamination, soil contamination etc., and criteria for selection of primary objectives of impact concerning groundwater, surface water, soils, air.

- The calculated score determines whether the second step will be applied (only for groundwater and soils). After the careful study of the available information about the groundwater and soils there is composing “Plan about additional investigations of the enterprise’s site” on the basis of the “Methods of ecological risk assessment”, the “Dutch target and intervention values for soil and groundwater” [Dutch target..., 1993], and the experience of other countries (Czech Republic) on these problems. The aim of the additional investigations is to evaluate soils and groundwater conditions and impact factors (polluters), in order to determine the extent of soil and groundwater contamination. The additional researches supplement necessary information about these evaluations.

There is approved a local monitoring network and the remedial measures. The a.m. Methods and Methodological Instructions have been applied for many industrial sites and reports are available in MoEW. Another methodology was developed under Twining project with Germany (the State Ministry of Environment and Agriculture – Saxony) giving rise to the ” Instructions for assessment and treating of previous damages” (2001) - It is based on Saxon methodology; - It includes quantitative assessment for groundwater; - The risk assessment was performed on 4 levels; - It uses the Computer Programme GEFA. The assessed factors (criteria) are: - Characterisation of the potential dangerous substances; - Characterisation of the emission of dangerous substances from the pollution

sources; - Characterisation of the discharge to groundwater; - Characterisation of the migration processes and behaviour of dangerous

substances in groundwater; - Characterisation of the present groundwater use and potential use of

groundwater. The a.m. instructions are applied in practice (2001-2002) for risk assessment of the bigger municipal waste landfills in Bulgaria – 59 in number. The number population (citizens) are disposing upon these waste landfills was then 6,596,170.

France The chosen methodology is pragmatic in order to be adapted on the different cases. In fact, the amounts of available data are strongly unequal between the groundwater bodies. This methodology is based on the analysis of: - data provided by existing monitoring networks (quality and quantity): comparison

with thresholds and limiting trend evolutions; - current and estimate pressures information; - vulnerability of the GWB. Otherwise, the spatial distribution of the points of measurement and their representativeness referring to the intrinsic vulnerability are taken into account. In case of lack of representativeness, it will be evaluated by expert judgement. In spite of the delay taken for the groundwater directive’s publication, a few hypotheses have been set up to assess the risk of failing good chemical status: - Point sources of pollution, particularly the industrial caused pollution, are in

general considered to be under control. The assessment is thus mainly based on diffuse pollution (nitrate and pesticide) and to a minor degree on other pollutants (essentially chlorinated solvents, chlorides, sulphate and ammonium);



- While waiting for the specifications of the daughter directive, a “good status” water is a water for which at every point concentrations are below the norms mentioned in the Drinking Water Directive;

- a risk is established in cases where concentrations: . exceed 80% of the parametric values mentioned in the Drinking Water

Directive (except for pesticides for which the 0.1µg/l limit has been kept, as well as for other pollutants like ammonium or chlorinated solvents),

. show significant and durable rise of the concentration of a given pollutant (5mg/l in 5 years for nitrates, 10mg/l in 5 years for sulphates, etc.).

Germany Referring to the received document Workshop LAWA-EUF Bonn, the following remarks can be made: Each German state has its own approach, however the methodologies used are in general comparable. Qualitative and quantitative risk assessments are divided respectively in diffuse and point pollution sources in groundwater balance and ecosystem issues. Concerning the diffuse pollution sources, the majority of the German states are using the 25mg/l nitrate concentration value over 33% of the surface of the groundwater bodies to define the risk criteria. However, other parameters and surface value are mentioned but with more variations between the states. Refuse Dumps associated with an impact over 33% of the surface of the concerned groundwater bodies are the generally mentioned cases of risk coming from point pollution sources. There are also here some variations of criteria between the states. Quantitative risks associated with water abstraction are in majority defined by the limit of variable % of withdrawal (10, 20 or 50%, depending on the state) compared to the recharge of the groundwater bodies. There is no common methodology concerning the risk of impact on ecological systems. Some states are not including ecological systems in the risk assessment. Others are in an ongoing evaluation. In general, the results show that for a majority of Groundwater bodies risks associated with diffuse pollutions.

Hungary The risk assessment is linked with three interactions: - Wetland ecosystem of river and lake; - Terrestrial ecosystem; - Man (and the products consumed by man. The risk assessment includes the following single assessment issues: - Groundwater bodies have been analysed with respect to diffuse nitrate-

pollution in detail, based on the evaluation of monitoring data (imm ission approach) and on estimated load to groundwater (emission approach).

- Pesticide could be analysed at country level only, because information is not enough for doing it for each groundwater body.

- Impact of point sources has been checked also from both immission and emission point of view. Taking the monitoring data of further 25 elements (NO2, Na, PO4, F, Al, B, Ba, Cd, Cr, Cu, Hg, Mo, Ni, Pb, Sb, Se, Sn, Zn, TPH, naftaline, fenol, BTEX, PAH, chlorinated hydrocarbons, halogene aliphatic hydrocarbons) we found 10 water bodies where in more than 20% of the wells the average concentration exceeded the Hungarian standard.

- The potential pressure from point sources based on a generalized hazard index lead to identify two further water bodies as “possibly at risk”.

- Higher concentration than the standard value in the case of As, NH4, Fe, Mn and organic matter content is considered as results of natural geochemical phenomena

The Netherlands

The approach for assessment of the chemical status of the GWB was the one used by the WFD but was also based on expert judgement. A GWB has a good chemical status when concentrations of polluting components:



- do not show any effect of saline of other intrusions; - do not reach the threshold values as defined in the WFD (thus only yet for nitrates

and pesticides); - are not as such that the environmental objectives for related surface water bodies

are not attained, the ecological status or chemical quality of the waterbody significantly decreased, or terrestrial ecosystem (directly dependent from the GWB) were significantly damaged.

Many groundwater dependant ecosystems are damaged because of ‘verdroging’ (drying out), a Dutch term that encompasses changes in the abiotic conditions of the ecosystem, i.e., changes in groundwater level as well as chemical composition of pore water.

Poland Expert judgement was used for assessment of risk at failing good chemical status (based mainly on existing inventory of pollutant sources, monitoring data and expert judgement.

Spain Criteria used to identify groundwater bodies according to the risk of failing good status are the following: - Bodies with proven pressure (with respect both to quality and quantity) where the

impact has been verified. Risk is sure and the water body must be declared in “Bad status”, and a further characterization must be done;

- Bodies with proven pressure, in which the impact has not been detected, though likely it will occur in the future. Risk is being studied and the water body must be declared a “Under risk”. In this case, monitoring networks shall be designed and put into operation in order to verify if the impact occurs or not;

- Bodies with a proven impact, but the pressure that originates it is unknown. Risk shall be studied and the water body must be declared as “Under risk”. A further characterization must be done;

- Bodies without significant pressure and with no impact. Risk does not exist and the water body must be declared as in “Good status”;

- Bodies without information about pressures or impacts. Risk cannot be assessed and the water body must be declared “Without data”.

The recipients considered have been: dependent ecosystems, water for urban supply, and groundwater.


The applied parameters for the risk assessment are a mixture of land use issues (urban, mining), identified contaminants (nitrate, pesticides, phosphate) and special cases (saline intrusions, point sources). A separate analysis is realised for each parameters. All the assessments aimed to determine failure of environmental objectives which is in fact wider than simply good status. This is the reason that trends assessment was included and also consideration of protected area objectives, where appropriate. As good chemical and ecological status has yet to be determined for surface water bodies, this was considered as an initial assessment to determine whether there may be pressures acting on the groundwater body that may impact the groundwater body itself. An assessment can then be made to determine if the potential impact on the groundwater body could have a detrimental effect on surface water bodies, protected areas and groundwater dependent terrestrial ecosystems. The receptors considered in the assessment are the surface water bodies but pressures on (and concentrations in the groundwater body) have been used to identify any risk to surface water bodies. The only assessment that explicitly tried to calculate the concentration at the rivers was for the assessment for phosphate as the river was considered the only target (due to the low EQS), unlike the other assessments where drinking water abstractions could have been considered as a target. This assessment used both the type of river (calcareous, siliceous, organic) and the baseflow index of the catchment.

Belgium (Flanders)

A preliminary evaluation of the quantitative status of groundwater was obtained by studying time series of piezometric measurements of at least 10 years. If the GWB shows no trend (with or without seasonal variations) the GWB has a good



quantitative status. If a decreasing trend is observed in the measurements, the GWB will have bad quantitative status. A preliminary evaluation of the groundwater quality was performed for each groundwater body by means of the results of the measurements on the phreatic monitoring network, focusing on the nitrate concentration. The Flemish Environmental Administration (AMINAL) uses the rule that a groundwater body has a good qualitative status if at least 95% of the total number of monitoring points is not exceeding the standard for nitrate.

3.8.3. Preliminary evaluation of aquifers being at risk

There is no information in some cases (Austria, Estonia) and a more or less ongoing classification in other cases (Bulgaria, Denmark, France, Germany, Lithuania, Poland, Portugal, Belgium, Spain and England & Wales). Referring to the WFD, the groundwater bodies must be classified into only two classes: “at risk” and “not at risk”. However, four countries are introducing a third class: France (“at doubt”), Netherlands and Hungary (“possibly at risk”) and Scotland (“probably at risk”). The question remains open about Finland: the so-called Class III (“groundwater areas which need further studies to find out the suitability of the area for water supply”) might be interpreted also as an intermediate case. Concerning France, the water basins which belong to this third class show following characters: - Some information are available, but further studies are needed to clarify the status; - Few or no information are available, a complementary programme of measurement and

following interpretation should be established and accomplished. Not all basin-districts in France are using this third class. 3.9. Open questions Is the surface of the whole country supposed to be covered by Groundwater bodies? If not, what are the justifications? A key issue for the evaluation of the assessment of GWB is the water abstraction (drinking water supply, irrigation, industrial uses). However this issue can vary between the countries. Thus it would be interesting to know for example the percentage of drinking water which is provided by groundwater abstraction. What kinds of actions are planned for the intermediary classes of Groundwater bodies which are qualified “at doubt”? 3.10. Concluding remarks This synthesis of replies to questions related to delineation and characterization of GWBs is based on very heterogeneous information and would require further complementary details. However some general features can be already pointed out.

3.10.1. Definition of the terms It is necessary to clarify the terms used and their significance so that all contributors or participants will speak the same language. A certain number of terms employed are not defined in the WFD. This is particularly the case for the following terms:



- River Basin District and Sub-river Basin district; - Groundwater Body and Sub-groundwater Body; - etc.

3.10.2. Current state of RBDs and GWBs It would be useful to have a more precise current state of the number of RBD, sub-RBD, GWB, sub-GWB.

3.10.3. Typology of GWBs The information provided by the responses to the questionnaire does not allow a synthesis on the typologies used by the various states of EU to delimit GWBs. However, the majority of the states are referring to the limits of the hydrogeologic entities, which are themselves mainly based on the lithological characteristics of the aquifers. To define the GWBs typology, the countries are combining in different proportion purely hydrogeologic characteristics with characteristics of intrinsic vulnerability (standard of flow, presence or not of an argillaceous cover, etc.) and/or the human pressures which influence the aquifers. Under these conditions, the delineated GWBs do not have the same significance and/or homogeneity in quantitative and qualitative terms.

3.10.4. Numbers and dimensions of RBDs Information collected is partial. However, it would be useful to have a general map, of RBDs for the whole of EU, even subject to later modification. In France, the surface of the 13 RBD varies between 1,085km² (Martinique) and 156,915km² (the Loire, coastal Vendee and Brittany).

3.10.5. Percentage of surface of the country covered by the GWB According to countries', the GWB cover or not the totality of the surface of the territory: - In the majority of the countries as in France and Austria, the GWB cover the totality of the

territory. - In other countries the GWB cover only one small part of the territory: e.g. in Finland the

GWB cover only 4.1% of the territory.

3.10.6. Numbers and dimensions of the GWB The number and the size of GWBs are extremely variable: - within the same country, - between the countries, - depending to the type of aquifer.

In Finland, for example, GWBs of class I and II count 3,700 and have sizes from 1 to 2km² (maximum 100km²). GWBs of class III are totalising 3,188.

France, counts 559 GWBs. Their size lies generally between 1 and 61,021km², the general median being of 733km². GWBs of the alluvial type have a median size of 209km² whereas



those of the basement type have a median size of 1,246km² and those of the sedimentary type have a median size of 950 km².

3.10.7. Assessment of risk not to reach good status of environmental objectives in 2015

Information provided in the answers to the questionnaire does not clarify sufficiently the methodologies and assumptions that are used to make the analysis of the Risk of not reaching good environmental status (RoF) of GWBs. It is therefore impossible to synthesise them properly. The total RoF in 2015 is the combination of the quantitative RoF and the qualitative RoF linked with the various pollutants which affect the GWB. The quantitative RoF is difficult to apprehend because it supposes that we have the measurements allowing comparing the recharge of the aquifers with the water abstraction and with the contribution of the aquifers to the flows of the rivers. This information is seldom available. Here is a key issue of research because it points out the necessity of a holistic view of the resource including surface water and groundwater and their interactions. The qualitative RoF concerns the various pollutants which affect the GWB. The sources of pollution can vary from one country or region to another: in the agriculture dominated zones the diffuse pollution (nitrates and pesticides) is important, in other zones it can be specific industrial pollution (e.g.: classified installations, etc.) or regional (ex.: exploitation of the oil shales or the lignite in Estonia, the urban pollution and industrial pollutions of old iron mines in Lorraine in France). The different countries don’t assess this kind of pollution in the same way. Thus, the preliminary evaluation of GWBs being at risk in France is mainly based on diffuse pollution (mainly nitrates and pesticides) and to a lesser extent on some other pollutants. Point source pollutions like industrial pollution are regarded as under control because they are treated by environmental measurements within the framework of specific existing policies. Generally, the assessment of the RoF of GWBs poses the following problems for each pollutant: - Assumption on the limit values to be used while waiting for the publication of the Daughter

Directive Water; - Representativeness of measurements: Appreciation of measurements at a point of a

network. The evaluation is done on the basis of series of measurement compared to the chosen limit values. Is just one critical measurement enough to give a bad qualitative status of the point?

- Representativeness of points: Aggregation of the assessment on the different points of the network for an overall assessment of the GWB. Thus, the density and the spatial distribution of the points are not always reflecting the characteristics of the GWB.

The results of the RoF for each pollutant are then combined. The result of this risk analysis is directly a function of the number of analyzed pollutants, of the number of measurements, their reliability and their representativeness (methods of space aggregation of the data). Taking into account the great diversity of the characteristics of GWBs with regard to: - their size;



- their hydrogeologic characteristics; - their vulnerability; - the data of the network of measurement (number and representativeness). The methodology of the RoF can be only pragmatic. But in the same time, one should be conscious of the limited degree of confidence of the resulting assessment. This is the case of the methodology used in France: A large GWB shows sometimes significant shifting in qualitative and/or quantitative terms. Thus specific parts of GWBs have been identified as sectors “at risk” (in other countries they have been defined as “sub-groundwater bodies”) for which specific measures of management are defined. Otherwise, it is advisable to take into account the variation of characteristics to define representative networks of measurement. The principal hydrogeologic characteristics to take into account for the characterisation of the GWBs are: - the lithological nature of the aquifers (typology of the GWB based mainly on

hydrogeology); - the nature of the flows (unconfined and/or confined); - relations between the GWB and the terrestrial systems ecological (wetlands) or surface

water (river) dependent; - the degree of homogeneity of the hydrogeologic characteristics and/or occupation of the

grounds; - karstic characteristics; - the overexploitation of the coastal aquifers inducing a risk of saline intrusion in the fringe

littoral; - the degree of homogeneity of a GWB according to:

. whether there are only one or several aquifers superimposed, more or less inter-connected aquifers,

. or of the regrouping of disjointed horizontal sectors; - the evaluation of the recharge of the GWB necessary to compare the average rate of

renewal of the resource with abstractions; - the intrinsic vulnerability which indicates the sensitivity of an aquifer to diffuse pollution: it

uses various characteristics of topography of the soil, of the occupation of the ground and the characteristics of the ground and the unsaturated zone.

The evaluation of the risk not to reach good environmental status in 2015 depends: - of course on what one understands by "Good environmental status". There is a need to

specify the environmental status of the GWB. Good status is currently partially defined in the WFD and must be supplemented by the DF being prepared. In waiting for the DF, different EU made assumptions which are not inevitably identical from one country to the other. In the WFD the GWB can have only 2 statuses: Good or Bad. It is seen that several states including France defined a 3rd status "with Doubt" which covers the GWB for which some data exist but for which complementary studies must be made;

- on the data available (space distribution of the network of measurement, types of measured parameters, frequency and reliability of measurements);



- for each point of measurement one will have to appreciate its status according to fixed rules (list of the concerned pollutants, comparison between a threshold and/or variable threshold for these various pollutants, etc.). Currently these approaches are variable from one country to another;

- one will have to aggregate these specific data to define the overall the status of the GWB. Which is the global status of a GWB made up of several sub-GWB having different status? The methods of aggregation of the specific data are not currently defined. They vary between the countries and, due to a lack of sufficient data, the status of a GWB must often be evaluated by expert judgement.

Taking into account the strong heterogeneity of the approaches of the various countries, it appears necessary to better harmonize the approaches aiming at defining the GWB and their status. Under all circumstances the methods must remain very pragmatic in order to adapt, as well as possible, to the hydrogeologic situations of the various UE countries. 3.11. References Normand, M. (2002), Délimitation des masses d’eau souterraine. Note explicative sur les relations quantitatives nappe-rivière et nappe-zone humide. Note BRGM/EAU - Version 1 du 24/09/2002. Normand, M., Mardhel, V. (2002). Consignes pour le remplissage des tableaux de synthèse de l’identification et de la délimitation des masses d’eau souterraine. Note BRGM/EAU, 5p. Normand, M. (2003). Note de synthèse bibliographique sur les relations qualitatives nappe – rivière. Note BRGM/EAU - Version 2 du 30/07/2003. Normand, M., avec la collaboration de Chadourne, D. (MEDD/DE) et des hydrogéologues des Agences de l’EAU et des DIREN déléguées de bassin - (2003). Mise en œuvre de la DCE. Identification et délimitation des masses d’eau souterraine. Guide méthodologique. MEDD/DE. Rapport BRGM/RP-52266-FR, 45p., 17fig., 1tabl. Normand, M. avec la collaboration de Chadourne, D. (MEDD/DE) et des hydrogéologues des Agences de l’EAU et des DIREN déléguées de bassin (2003). Mise en œuvre de la DCE. Caractérisation initiale des masses d’eau souterraine. Guide méthodologique. MEDD/DE. Rapport MEDD/DE, 57p. MEDD Direction de l’Eau, Aquascop (2003). Mise en œuvre de la DCE. Identification des pressions et impacts. Guide méthodologique version 4.1, mars 2003. Common Implementation Strategy for the Water Framework Directive (2000/60/EC). Wetlands Horizontal Guidance. Horizontal Guidance Document on the Role of Wetlands in the Water. Framework Directive. Final Draft Version 8.0. 7th November, 2003. Loi n°2004-338 du 21 avril 2004 portant transposition de la directive 2000/60/CE du Parlement Européen et du Conseil du 23 octobre 2000 établissant un cadre pour une politique communautaire dans le domaine de l’eau ; JO n°95 du 22 avril 2004, p. 7327. Nilsson, S., Langaas, S., Hannerz, F. (2004). International River Basin District under the EU Water Framework Directive: Identification and planned cooperation. European Water Management online, Official publication of the European Water Association (EWA).



Collectif (2004). Common implementation strategy for the Water Framework Directive (2000/60/EC) – Groundwater body characterisation. Technical report on groundwater body characterisation issues as discussed at the workshop of 13th October 2003, 11 avril 2004. Collectif (2004). Common implementation strategy for the Water Framework Directive (2000/60/EC) – Groundwater risk assessment – Technical report on groundwater risk assessment issues as discussed at the workshop of 28th January 2004, 12 octobre 2004. Normand, M., Gravier, A. (2005). Caractérisation initiale : premiers éléments pour une synthèse nationale. Piste de réflexion pour la caractérisation détaillée. Diaporama - Note BRGM n° 15-05 ; réunion du groupe de travail « Référentiels masses d’eau souterraine et. BD RHF V2 » du 22 mars 2005. Normand, M., Gravier, A. (2005). Éléments de synthèse complémentaires sur la caractérisation initiale (RNABE en 2015) – Pistes de réflexion pour la caractérisation plus détaillée – Note BRGM/EAU n°88-05. Présenté lors de la réunion du groupe référentiels masses d’eau souterraine et BDRHF V2 » du 13 juin 2005. Mardhel, V., Normand, M., Gravier, A. (2005). Mise en œuvre de la DCE. Référentiel cartographique national des masses d’eau souterraine (version 1). Rapport final BRGM/RP-53923-FR, mai 2005. Normand, M., Gravier, A. (2005). Mise en œuvre de la DCE. Premières synthèses des caractéristiques principales et secondaires des masses d'eau souterraine et de l'analyse du risque de non atteinte du bon état environnemental en 2015. – Pistes de réflexion pour la caractérisation plus détaillée. Rapport BRGM/RP-53924-FR, août 2005. Arrêté du 16 mai 2005 portant délimitation des bassins ou groupements de bassins en vue de l’élaboration et de la mise à jour des schémas directeurs d’aménagement et de gestion des eaux. JO n° 113 du 17 mai 2005, p. 8556, n° 23.



Appendix 3.1

Statistical characteristics from the participant countries and their Groundwater Bodies

We have synthesised in a table form some characteristics according to the responses on the questionnaires and to some bibliographic references. In this table we have presented following information: - Name of countries and occasionally the region (Belgium and United Kingdom); - Number of Water Basin District; - Country or region surfaces; - Population; - Number of GWB; - Average surface of GWB; - Range and median of GWB; - A last column “specific point” emphasizing some particularities. Remark The average surface of GWB gives the result of the whole surface of GWBs divided by the country surface. It is just an indicative value. Thus this calculation procedure is correct only if all GWB are shallow GWB covering the whole countries. In France the GWBs are covering the whole country surface, but the deeper, covered GWBs are added. Thus, the country surface is 646,321km², but the total surface of GWBs is 1,094,514km². Otherwise, in Finland, only 4,1% of the country surface is related to the identified GWBs. We have also expected to provide further information about other parameters but the available knowledge at that work stop didn’t allow it: - Number of points for groundwater monitoring; - Percentage of groundwater withdraw in the drinking water balance; - Systematic typology of GWBs.

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Countries Number of

Water Basin District

Area of the

country (km²)

Area of the WBD

(km²)

Population (106 inhab.)

Date of réception question-

naire

Number of GWB

Average surface of

GWB (Km2)

Surface of GWBs,

Range and Median (Km²)2

Specific points

Germany 16 102 356718 82,4 14/04/2005 1200 297

3 classes of vulnerability; definition of pressure from diffuse pollution sources on a state level based on model calculation, no uniform approach and methodology for risk assessment The main pollution source is diffuse pollution and it related to agriculture: nitrate, pesticides Detailed and exhaustive analysis of background quality of aquifers (separation of natural and influenced groundwater component)

Austria 31 33 83859 8,1 11/04/2005 135

1382 621 6082

Single GWB: 7 – 1200 Groups of GWBs: 9 –

9600

The following characterisation of groundwater bodies was applied: single GWB – groups of GWBs; shallow GWBs - deep GWBs; predominantly porous media - karst - fractured GWBs (and subtypes)

Belgium (total) 30518 10,3

Belgium (Flanders) 21 13428 5,95 11/04/2005 42 320

In practice the underground of Flanders has been first divided into groundwater systems, who have later been divided into groundwater bodies. The subdivision of the groundwater systems into groundwater bodies has been done horizontally and vertically, so the groundwater bodies are superimposed in the depth ANNEXES 1, 2 and 3 must be provided!!!

Belgium (Walloon & Brussels)

No information 16785 3,40 11/04/2005 33 509

Hydrographical limits were also used for fitting to the districts as well as sometimes to sub-basin (especially where appropriate for international management) Excepted for deep aquifers, every GWB is supposed to have an influence on surface water and as a consequence on surface water or terrestrial associated ecosystems Agriculture and urban management (sewage), are the main risks identified for groundwater. They must be considered as a diffuse source of pollution. Main pollutants are nitrates and pesticides

2 Referring to: ”International River Basin Districts under the EU Water Framework Directive: Identification and planned cooperation”, S;Nilson, S.Langaas and F. Hannerz; European Management

online – Official publication of the European Water Association (EWA); 2004 3 Referring to: Groundwater body characterisation, April 2004

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Countries Number of


Area of the

country (km²)

Area of the WBD

(km²)



naire

Number of GWB

Average surface of

GWB (Km2)

Surface of GWBs,


Specific points

Bulgaria 3 111000 7,9 11/04/2005 170 653

2D system is created but with 8 GIS layers: the dividing of groundwater bodies in layers in GIS was made on the basis mainly of the geological age of the water bearing rocks or deposits (See annex 2) Only for some cases is clear now that these protected areas with habitats and species are directly dependant from groundwater The percentage for different types of land use are determined by GIS and input in data base for groundwater – for 170 GWB in ExEA Risk Assessment of GWB is partly based on a point systems linked with the different pressures; otherwise a German Risk Assessment methodology has been applied for the biggest municipal waste landfills (59)

Cyprus 9250 0,85

Denmark 12 131 122

43080 5,3 11/04/2005 500 - 600 72 - 86

Following three parameters have been used for the characterisation of the groundwater bodies: Lithology - Contact with surface water - Redox conditions There is also a detailed typology of groundwater - surface water interaction according to the dominant water input: Precipitation dominant wetlands - Groundwater dominant wetlands - Surface water dominant wetlands National analysis of vulnerability related to nitrate has been made

Spain 151 504790 41,1 15/04/2005 504 1002

In the pilot river basin Jucar: 48 –

7421

The pressure from diffuse and point source of pollution has been evaluated by the percentage of land used for different activities GWB are defined "under risk" or "good status" , but also a third class has been introduced: "without data" Vulnerability has been evaluated using the "COP method": C refers to the surface conditions that have influence in the water flow to the maximum infiltration zones, where the possibilities for human actions are less. Factor P refers to the influence of rainfall for the transport of pollutants to the saturated zone. Factors C and P are corrective factors for the protection degree of aquifers defined in factor O.

Estonia 11 45227 1,4 11/04/2005 15 3015

There are 8 regional aquifers (aquifer systems) in Estonia, which were divided into smaller units according to surface water catchment’s areas Groundwater - surface water interaction study has been performed in mining area in north-eastern Estonia

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Countries Number of


Area of the

country (km²)

Area of the WBD

(km²) Population (106 inhab.)


naire

Number of GWB

Average surface of

GWB (Km2)

Surface of GWBs,


Specific points

Finland 81 338147 5,2 02/08/2005 3700 (to

be grouped)

Classes I & II: 1 to 2

Class III up to 3188

Aquifers have been also classified on the basis of suitability for water supply and the need of protection into three classes: Class I holds areas important for water supply - Class II holds area suitable for water supply - Class III holds other groundwater areas, which need further studies to find out the suitability of the area for water supply The GWBs are not covering the whole country surface (only 14100 of 338147 km2: 4,1%). The most frequent surfaces of GWB lay between 1 to 2 km2. The biggest is about 100 km2.

France (total)

Most Frequent: 1 –

2 Greatest:

approx. 100

The whole country surface is covered by GWBs

France (metropole) 9 543965 61,2 15/02/2005 533 1021

Proven point sources pollution (urban and industrial) are at first regarded as under control (consequences?) Detailed algorithm for the risk assessment analysis based of 80% or the value threshold and the point representativity

France (overseas

territories)4 41 88932 15/02/2005 25 3557

Greece 131626 11

Hungary 11 93030 10,2 FEB 2006 108 861

More than the half of groundwater bodies are cross-border. Basic division in cold waters and thermal waters Identification of 33 type of biotopes linked with groundwater bodies Pressure estimation of point source pollution using a “generalized hazard index”

Ireland 68895 3,8 972 532 5 - 1400 Italia 301316 56,4 No answer

Latvia 64589 2,4

Lithuania 41 65301 3,7 03/08/2005 16 sub-GWBs

62 108842 3,7 – 20,2

12 of a total of 16 GW sub-bodies are at potential risk Main quality problem reflected by the monitoring: SO4 and Cl (mineral water intrusion due to water use), NO3 (urban impact)

Luxembourg 2586 0,4 Malta 316 0,39

4 Decree from the 16th May 2005 concerning the delineation of basins or groups of basins in order to elaborate and renew the Water Master Plans. 14 basins have been identified including the river

basins of the island of Mayotte. This last basin is not included in the French reporting of 2005.

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Area of the

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Area of the WBD

(km²)



naire

Number of GWB

Average surface of

GWB (Km2)

Surface of GWBs,


Specific points

Netherlands 4 42 41029 15,9 11/04/2005 374 110

The classification was done on 2 levels, national and regional. The first category represents the GWB covering the whole river basin district (20 GWB’s; category 1), whereas the second represents the groundwater resources used for human consumption (374 GWB’s; category 2). GWB are defined at risk (129) , but also an intermediate class has been introduced: possibly at risk (200) Risk analysis is only yet based on nitrate and pesticide

Poland 312690 38,7 11/04/2005 59 5300

The GWB has been defined only for freshwater systems (geothermal and mineral water systems have not been considered) Problem of water quality due to lowering of water level: oxidation of pyrite, consequences from mining activities Information layers of GW endangering (five classes) existing at electronic version of Detailed Hydrogeological Maps (1:50000) covering whole country Main quality problem: Fe and Mn (only background?), NO3 and N02

Portugal 91906 10,8 août-05 58 632 1585 5,1 – 54,8 GWBs surface from 1 to 54,8 km2

Czech Republic 78870 10,3

United Kingdom 241751 59

United Kingdom

(England & Wales)

11 (11, 126 CAMS5)

162 962 53,5 25/04/2005 356

(500 – 800)

458 (326 – 204)

Pragmatical characterisation of the groundwater bodies in 4 types referring to the Catchment’s Abstraction Management Strategy (principal aquifers, secondary aquifers, unproductive strata, significant drift aquifer) The results of the exposure pressure for groundwater abstraction as a percentage of river water body (natural flow exceeded 70% of time) for different river ecology are divided in to 5 sensitivity bands: Very High to High, to Moderate, to Low and down to Very Low sensitivity bands; for lake: High, Moderate, Low and No Exhaustive analysis of all pressure, then the GIS layers for all of the pressures were superimposed to establish the total risk to each of the groundwater bodies The vulnerability is assessed by considering following parameters: Soils (High, Intermediate and Low vulnerability), Aquifers (Major, Minor and Non aquifers), Source Protection Zones (zone I, II and III) and Unsaturated Zone (depth)

5 CPAM: Catchment Abstraction Management Strategies

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Countries Number of


Area of the

country (km²)

Area of the WBD

(km²) Population (106 inhab.)


naire

Number of GWB

Average surface of

GWB (Km2)

Surface of GWBs,


Specific points

United Kingdom

(Scotland)

2 (2) 78789 5,5 11/04/2005

124 (Approx.

100) 635

Typology of GWB based on following parameters: overlying strata, transmissivity, confinement, porosity, GW chemistry, depth, flow path / GWB are defined at risk (26), but also an intermediate class has been introduced: possibly at risk (11) Chemical problems in regard of water abstraction: seawater and minewater Main quality problems are nitrates, phosphates and pesticides

Slovakia 49040 5,4 Slovenia 20250 2 Sweden 410934 8,9 TOTAL



Chapter 4

Concepts for characterisation of aquifer regarding transport and fate of pollutants

Prepared by H. Pauwels (BRGM), Z. Simonffly (BME), R. Kunkel (Fz-Jülich)

on behalf of all WP2 contributors



4.1. Generalities Main elements and parts of a sub-surface flow system within an aquifer through surface waters and terrestrial dependant ecosystem are illustrated through figure 4.1. The recharge from precipitation or from surface waters supplying the groundwater system can be polluted by surface originated pollution and this pollution front is moving towards the location of discharge (wells, springs, surface waters or terrestrial ecosystems). There are local, intermediary and regional type flow systems, with very different travel time between the recharge and the discharge zones (from a few months to several thousand years). Along the pathway the concentration of pollutant can decrease due to the mixing of waters from different origins and natural attenuation of the pollutant mass due to (bio)geochemical processes. In certain parts of the aquifer, the concentration of certain substances can be influenced by natural geochemical phenomena resulting in high natural background level. To be noted, that the groundwater abstractions can considerably change the natural recharge/discharge conditions and as a consequence the concentration of pollutants.

local recharge local rechargelocal recharge

regional recharge

dischargedischarge

zone of influence of anthropogenic surface originated pollution

regional geochemical background contamination

aquiclude

Gro

undw

ater

body

local geochemical background contamination

Figure 4.1 - Main elements of a subsurface flow system.

The description of the main characteristics/parameters relevant with respect to pollutants behaviour along the above described pathway as well as the methods used to assess pollutants behaviour is the purpose of the following chapters. However, before tackling that question and to make usable in a simple way the provided information for application of the methodology to derive environmental threshold at specific cases, a typology of aquifer has to be defined. Consequently, for each type of aquifer, specific characteristics relevant with respect to pollutant behaviour will be defined. In the same way, the basic principle of pollutant attenuation is described. 4.2. Typology for hydrogeochemical characterisation of aquifers As previously highlighted, Member states directly or indirectly (through delineation of groundwater bodies) have applied different classifications of the aquifers. According to the countries, the two most important classification criteria are the following:



- Water abstraction based classification (England and Wales, Finland); - Hydrogeological based classification (the others countries). In the case of the hydrogeological based classification, it appeared that criteria are a mixture in various proportions of purely hydrogeologic characteristics with characteristics of intrinsic vulnerability (standard of flow, presence or not of an argillaceous cover, etc.) and/or the anthropogenic pressures which influence the aquifers. Harmonisation of the national approaches - as in the case of surface water typology – would have lead to a simple and robust typology at European level. Despite the size of the task, such an harmonisation would not have been necessarily appropriate to the purpose of threshold definition, and in any case local adaptations would have to be applied. During the WP2 Workshop, partners discussed and agreed on a typology adapted to the definition in environmental thresholds for which attenuation processes (extinction/alteration or dilution) can be taken into account. The prime criteria of proposed typology for hydrogeochemical characterization include: - Lithology:

. Limestone (karstic/non karstic),

. Chalk,

. Volcanic rocks,

. Crystalline basement,

. Shales and schists,

. Sands and gravels,

. Sandstones,

. Evaporites,

. Clays, marls; - Saline influence. The further criteria will enable to supplement and precise aquifer description to comprehend pollutant behaviour. - Hydrodynamics (recharge, residence time, topography, leakage…); - Redox conditions; - Particular occurrences (e.g. Organic matter, Oxides, Sulphide minerals…); - Geological age. Those parameters are considered as determining for the Natural Background Level of any chemical element. The knowledge of Natural Background Level (NBL) is a key point of threshold determination as concentration lower than NBL value could not be reasonably reached. The relationships between aquifer typology and NBL is the subject of chapter 6. The typology represents also a key point regarding the behaviour of any chemical elements introduced within the water consequently to human activity. This relationship is the purpose of Chapter 5. This definition of such a typology does not signify that adaptation and addition of other descriptors specific to the some local conditions are to be dismissed.



Figure 4.2 – Map of hydrogeochemical units (typology for hydrogeochemical characterisation)

applied to first aquifer compartment within several European countries. 4.3. Characterisation of attenuation between source of pollution and receptor The concentration of the pollutants is modified along the pathway from the source toward the locations of considered receptor (figure 4.3). At any point located before receptor, all processes able to modify the concentration before reaching the receptor have to be taken into account. Actually, these processes can be modelled in rather simple way. Attenuation of the pollution between two points A and B along a flowpath in the aquifer can be defined as the ratio of concentrations in A and B: CB/CA with CB<CA. Actually this ratio depends on both dilution processes and those processes impacting the pollutant mass. Therefore, the attenuation can be simply formulated as:

CB/CA = (1- δaq).βaq

where δaq is the index of pollutant mass decrease and the βaq index of dilution, with δaq= 0 when the mass of pollutant is totally preserved and δaq= 1 when pollutant is totally removed from groundwater. βaq= 1 when dilution of pollutant is negligeable and decrease up to 0 with increasing importance of diluting processes.



r

Figure 4.3 - Simplified schema of pollution endangering receptors. For the description of the evolution of the concentration of a pollutant between a point B located within an aquifer and point C located in a river or a terrestrial ecosystem, beside the groundwater, surface runoff or recent infiltration contributes as well to the water supply of the ecosystems and concentration of the pollutant in these water can be far from negligeable. Additionally, geochemical processes can reduce the pollutant mass in the hyporheic or in the root zone, thus the impact of the mixing (βex = index of mixing between Groundwater water and the other water and, Cex= concentration of the pollutant in the other water) and the index of decrease of pollutants in the terrestrial or surface water ecosystem (δr) should also be taken into account. Therefore, assuming geochemical processes occur after waters mixing, the concentration of the pollutant at point C can be expressed as following:

CC = [(1-βex).CB + βex.Cex].(1-δr)

where βex is the mixing index. By replacing CB into this equation by its expression from previous equation, it is possible to rely concentration of a pollutant in a surface or Terrestrial ecosystem with its concentration at any location in the contributing groundwater. Such equations can be a priori used to derive thresholds. For example, if we consider that the objective of concentration Cc is known, Threshold concentration (Cth ) to be applied at point A would be described by the following equation:

Cth ≈ [Cs – (1- δr).βex.Cex]/[(1- δr).(1-δaq).(1-βex)]. Estimating these indexes constitutes a real challenge in most of the settings. Although robust estimate are expected locally, necessary information is generally not available. Nevertheless,

Crec

root zone, lakes, rivers,

Crec,aq

mixing

diminuation and

mixing

aquatic or terrestrial ecosystems

receptors

human

Cex

Crec = [(1-βex).Crec,aq + βex.Cex].(1-δr) Crec,aq = (1- δaq).βaq.Ci δaq: index of decrease of pollutant in the aquifer βaq: index of mixing in the aquifer δr: index of decrease outside the aquifer βex: index of mixing from external source

source of pollution

polluted recharge

aquifer

Ci Crec,aq

mixing

diminuation and

mixing


receptors

human

Cex


source of pollution

polluted recharge

aquifer

Ci Crec,aq

mixing

diminuation and

mixing


receptors

human

Cex


source of pollution

polluted recharge

aquifer

Ci



through application of some methods, it remains possible to acquire this information. For the present work, it is also considered that typology of the aquifer is a key point regarding the possibilities of attenuation including both dilution and decrease of pollutant mass. Chapters 5, 7 and 8 deal with evidence of attenuation in different settings and methodologies to be applied for attenuation assessment in the aquifer, at the interface between groundwater and surface water and at the interface between groundwater and terrestrial ecosystem respectively. Finally conclusions regarding behaviour of pollutants in aquifer and at the interface with surface water and terrestrial dependent ecosystem can be largely modified by different actions on the water quantitative status. It is the purpose of chapter 9.



Chapter 5

Hydrogeological processes

Philippe Meus, L. Piron (ULG) Hélène Pauwels (BRGM)

Zoltán Simonffy (BME) Manuela Ruisi, Paolo Traversa, Alfredo Di Domenicantonio (Abteverre)

Kristine Walraevens, Marleen Coetsiers, Marc Van Camp (LAGH-UGENT) Frank Wendland (FZ-Jülich)

Johann-Gerhard Fritsche (HLUG) Rüdiger Wolter (UBA-D)

Andres Marandi (UT) Carlos O. Miraldo, Clara Sena, M.T. Condesso de Melo (UNI-AVEIRO) Carlos Martínez Navarrete, José Antonio de la Orden Gómez (IGME)

Jan Hookey (EA), Gorova Rossitza (EEA)

With the contribution from

David Allen, Marianne Stuart (BGS)



The role of aquifer settings in the natural groundwater composition, referred to as natural background levels, is mentioned in the chapter 4 of the compendium report of Workpackage 1 and is more detailed as well in the chapter 6 of the present report. Anyway, these descriptions are only considering naturally, geochemically occuring substances. The present chapter is an attempt to generalise the fate and particularly the potential “natural” attenuation of any pollutant introduced in groundwater from an anthropogenic source. It may act to decrease the amount, the concentration or the flux of the pollutant on its travel. It deals with processes related exclusively to the water movement (flow processes) as well as processes involving physico-chemical reactions6 (physico-chemical processes). A classification of these processes is given in Carey et al. (2000). Some of these processes (i.e. geochemical processes) may be the same as those that allowed groundwater to achieve its background but they are not necessarily acting at the same scale of space and time. The following considerations are implicitly based on the principle source-pathway-receptor but: - they do only consider the attenuation inside the aquifer, from an intrinsically point of view,

and not the source itself, nor the receptor; - they do not refer to any specific space and time values, but only to potential processes,

anywhere, at any time. This point is left to the appreciation of the user who must additionally consider from which level its monitoring network is indeed representative7.

One must keep in mind that the limit between a naturally occurring substance, and a widespread occurrence of the same pollutant anthropogenically introduced, may sometimes be very fuzzy8. The present chapter thus focuses on the potential attenuation that should substantially reduce the level of the pollutants, especially in order to avoid taking unnecessary measures. Since both flow and physico-chemical processes are strongly controlled by the intrinsic nature of the aquifer, a good conceptual understanding of pollutant dispersion and persistency or attenuation can be based, under certain assumptions, on a simplified typology of aquifers that highlights the influence of these processes. The principle is very similar to what will be suggested for determining natural background. In order to approach a European representativeness , each participant was asked to contribute in the description of attenuation characteristics for several typical aquifers present in their country. The results presented here use both a state-of-the-art knowledge and this collected information. These results will not refer specifically to each individual pollutant, except in some case studies. It is preferentially based on the search for the main factors controlling the occurence of the processes in one given aquifer. The link may be done further by crossing this information with that on the processes affecting pollutants (cf. WP1 section).

6 When speaking about reactions one should also consider the transformation of one contaminant into another (metabolites). 7 Some are using “alert” networks close to the source of the pollution while others use networks closer to the receptors. 8 Note that in both cases, any measure to reduce the level of pollution may also become tedious and economically unrealistic.



5.1. Importance of aquifers in the context of the WFD In a general manner, the WFD is dealing with the quality (and the quantity) of all water that takes part in the hydrologic cycle and may accumulate sufficiently inside the geosphere to provide a resource for living organisms. It considers the impacts that may result (or may have resulted) from any anthropogenic activity and intends to prevent or control those activities in order to progressively improve the status of water. Subsurface water is the part of that water that can be found beneath the surface (Bear & Verruyt). Hydrogeology is the science that deals with subsurface water. Many authors, and the WFD does so, introduce a restriction by defining groundwater as only the subsurface water that accumulates in a saturated zone. As far as the majority of the sources of pollutants are entering the subsurface through unsaturated or temporarily saturated layers, many of the processes described in the present chapter will also be relevant to the unsaturated zone and must obviously be taken into account in assessing the attenuation of pollutants along the pathway from the source to the receptor. Anyway, those effects in the unsaturated zone may become so complex, site specific and variable that they can not be used in a pragmatical approach of groundwater thresholds at the groundwater body scale. They will not be dealt with in this chapter and rather own to be used in assessing the chemical status at each specific monitoring point. Those aspects dealing with the attenuation of pollutants in the unsaturated zone are sometimes referred to as vulnerability characteristics. The proposed methodology has to tackle with a second obstacle in its conceptualization. It arises from the possible insuitability between aquifers and groundwater bodies. The aquifer may be defined (Bear & Verruyt) as a geological formation which (i) contains water and (ii) permits significant amounts of water to move through it under ordinary field conditions. De Marsily (1986) adds a “resource” nuance by defining an aquifer as “a layer, formation, or group of formations of permeable rocks, saturated with water and with a degree of permeability that allows economically profitable amounts of water to be withdrawn”. The WFD is using a similar definition: ... but sets the quantitative limit for the water potentially withdrawn to 10m3/h. All these definitions require implicitly that significant hydraulic connections may exist throughout the aquifer. It results that one single aquifer is not necessarily restricted to one lithological or litho-stratigraphic unit and that this entity is more corresponding to a notion of system. However the most important point is that the connections can be considered as relative and depend on the scale at which they are considered. On another hand, groundwater bodies have been defined in the WFD as management (indeed practical) units that may include one or several aquifers. They can also be delineated according to their status or to non hydrogeological criteria. As a consequence, Member States have defined their own groundwater bodies in a very heterogeneous way, considering different vertical and horizontal scales. For this reason, setting threshold values at the groundwater body scale may lead to some trouble when assessing the status of groundwater bodies that have been designed including several contrasting aquifers. In this case it is recommended to address each aquifer in a separate way through distinct monitoring networks and distinct status assessments.



Flow processes Physico-chemical processes (pollutant specific)

Attenuation

Residence time Kinetics

5.2. Concept of the aquifer control on pollutants The fate and particularly the attenuation of pollutants is mainly controlled by the characteristics of water flow. Flow characteristics directly determine processes such as dilution or dispersion, while, through its control on the water velocity and the residence time, it also obviously modulates the physico-chemical processes. This dual effect is shown in the following diagram (figure 5.1). Since general flow characteristics may be estimated for each type of aquifer, it should be possible to establish a list of potential processes involved in each kind of aquifer. Unfortunately, the situation is somewhat complicated because flow or physico-chemical processes may be influenced in different ways for a same aquifer under different conditions. These conditions may result from the combination of several specific controls, either external or internal (intrinsic), on the aquifer. From the internal point of view especially, groundwater are not often homogeneous over distances of any practical significance. This situation is illustrated in figure 5.2. A suggested list of these controls can be found in the synthesis at the end of the chapter.

Figure 5.1 - Dual effect of flow and physico-chemical processes on pollutant attenuation.

Figure 5.2 - Modulation of pollutant attenuation through external and internal controls on a typical aquifer.

Main aquifer type

Pollutant input in the aquifer

Pollutant output to the receptor

Attenuation Internal controls

External controls



5.2.1. General characteristics influencing both the flow system and physico-chemical processes

5.2.1.1. Saturated versus unsaturated (vadose) zone

The attenuation of pollutants through the unsaturated zone may appear to be neglected in groundwater assessment for the WFD because monitoring points are restricted below the water table. Since pollutions may both strongly decrease and be delayed in this zone, it would be a nonsense to simply forget it. That is the reason why its specificities are shortly described here. In the unsaturated zone, vertical flow is predominant, thus highly dependant on the vertical permeability which may vary with the water saturation and according to macro-heterogeneities. Several overlying layers are usually distinguished. In the case of a fissured aquifer for instance the following layers can be found (adapted from Cost 620): - topsoil (biologically active zone); - subsoil (non lithified material); - unsaturated overlying formations; - unsaturated part of the water bearing fissured formation. Layers with lower permeabilities can also sometimes create perched, saturated or temporarily saturated, levels. Flow features may especially become complex where a non aqueous phase liquid (NAPL) is present. The interface between unsaturated and saturated zone is particularly important, especially where an important water table fluctuation exists. It is the place where some processes like for instance oxidation are predominant.

5.2.1.2. Redox conditions Redox conditions are of prime importance regarding degradability of pollutants. Some of them are sensible to oxidizing conditions- degradation occurs through reaction with oxygen. Other are degraded under reducing conditions. Pollutant properties regarding redox conditions are described within WP1 report. Here, the present report deals only with redoc conditions within aquifers. Generally at shallow depth, oxidizing conditions prevail, but they are also many examples where reducing condition exists just below the groundwater level. Either chemical composition of recharge water or special occurrence within the rock matrix can be involved in the development of reducing conditions: Occurrence of Organic matter or sulphide minerals may ply this role. The penetration of groundwater in some aquifers may become so deep or so distant from the recharge zone that a strong depletion of oxygen is observed and anoxic (reducing) conditions may install. This is often the case (figure 5.3) in thick aquifers or when the aquifer is dipping under an aquiclude or an aquitard (confined or semi-confined aquifer sensu stricto). This situation is essentially due to the fact that these parts of the aquifer contains



water with relatively high residence time and low turnover time9. The transfer of pressure can be nearly instantaneous while the movement of water is strongly restricted due to the absence of close outlets. Dilution processes are not very effective but they are compensated by the very weak pollutants entries and the sustained degradation of these pollutants, provided they are unstable under reducing conditions. Mixing with fossil marine waters or crustal waters is even sometimes reported, contributing to a more specific hydrochemical context. In the case of confined water, anthropogenic effects may of course be of significance for the resulting quality of withdrawn water (mobilization created by an upconing). Reducing conditions may also have some adverse effects on pollutants by the solubilization of sulfides and metals like Mn, Fe or As.

5.2.2. Residence time

5.2.2.1. Residence time in aquifers Aquifers may be considered as reactors transforming a pollution from its input towards an output to the receptors. These transformations can result from complex processes involved in several successive or parallel compartments or subsystems. A mean residence time of water in the aquifer, or at least an order of magnitude of it, can be used in a first approximation to assess the capacity of attenuation of the pollutants. The higher is the residence time, the higher is the chance that attenuation of the pollutants may be effective. The residence time of water is thus a good indicator of the potential attenuation of the pollutants10. The mean residence time of water can be approximated by dividing the volume of water contained in the aquifer by the discharge at the outlets, as follows:

T = V/Q. The estimation of residence times distribution in the aquifer system often requires the use of a flow model. Such models (plug flow, perfect mixer, dispersion, stagnant zones, recirculation…) are described in IAEA (1990). Appropriate information may be also provided by environmental (isotopic) tracers (see section 5.2.2.2) However the mean residence time of water is only a statistical value at the aquifer scale. For these reasons any value given in the literature can only be used as an order of magnitude. The table 5.1 gives a rough estimation for some typical aquifers.

Table 5.1 - Order of magnitude for water residence time in typical aquifers.

Aquifer type Order of magnitude for water residence time Karst 1 day to 100 years (depending on conduit or diffuse flow) Carbonate non-karst 1 to 100 years Sandstones 1 to 100 years (10,000 years for deepest aquifers) Sandstones, siltstones, claystones (clastic series) 1 to 100 years

9 Residence times higher than several hundredth of years are common in confined aquifers. 10 Note that the higher the residence time is, the slower the replenishment of the aquifer will be and the longer the recovery of

the groundwater will be in case of contamination.



Chalks 10 to 1,000 years (depending on fissured or diffuse flow) Alluvial gravels 5 years Sands and gravels 1 to 1000 years (depending on flow system scale) Crystalline (weathered) 10 to 100 years and more Volcanic 1 to 1,000 years (depending on depth)

It is worth pointing out that only rough estimation area presented on table 5.1 as in European Aquifer, Pleistocene water b( >10000 years) have been identified. Mean residence time values must also be used carefully for additional reasons: - due to the high complexity of the medium the distribution of residence times may be highly

heterogeneous. Dual porosity or multiple reservoirs effects may induce a more complex distribution. The reality of flow lines into the aquifer is much more complex due to the heterogeneities. All water is not necessarily participating to flow. A volume V representing only the circulating water should be used instead of the whole volume of the aquifer. The value also depends on a wide variety of factors controlling the aquifer functioning;

- the methods used to assess this mean value may also reveal uncertainties, or be representative of one particular component of the flow. Chemographs of major elements of springs or tracer tests are usually used for short residence time estimations while isotopic methods are used for older water estimations;

- a source of pollution does exceptionally concern the whole aquifer (it is more often a point-source pollution with regard to aquifer scale). The spreading of the pollutants depends on more local flow conditions. The effective residence time also depends on the scale of the involved flow systems;

- the residence time of water is hardly representative of the residence time of the pollutant itself. This latest can be much longer due to different effects (i.e. retardation, trapping…). A lot of pollutants do not behave like ideal11 solutes.

5.2.2.2. Dating groundwater with environmental tracers

Environmental tracers are defined as natural or anthropogenic compounds or isotopes, which variation in abundances in the near-surface environment can be used to determine timescale of environmental processes. Among the environmental tracers proposed by scientific community to assess groundwater mean residence time, only tritium is part of the water molecule and can thus actually“ date” the water. All other dating methods rely on dissolved compounds the abundance of which in groundwater depends on physicochemical and biological.

a) Detecting and dating modern groundwater By modern groundwater, we mean those recharged within the past few decades • Stable isotope There is a strong correlation between temperature and stable isotopes (δD, δ18O) in meteoric waters. After recharge, these seasonal variations generally disappear in the reservoir within 2 to 3 years. Very short mean residence times can thus be determined by monitoring the seasonal variations in δD or δ18O in precipitation and groundwater over at least 1 year.

11 "Ideal" means here that it strictly conforms to the water movement.



• Tritium Tritium (3H) is probably the most commonly employed radioisotope used to identify the presence of modern recharge. It is the only radioactive isotope of hydrogen, and has a half-life of 12.43 years. Atmospheric thermonuclear testing, in the 1950s and 1960s, caused large increases in the tritium content of precipitation worldwide, providing a useful environmental tracer for groundwater originating from this period. The final year of megaton tests (1962) generated a huge peak, which appeared in precipitation in the spring of 1963, Concentrations of 3H in precipitation are now largely back to natural, cosmogenic levels. Both qualitative and quantitative approaches to dating groundwaters with 3H have been proposed: They include: - Velocilty of the 1963 3H peak –Identification of the bomb-spike preserved in groundwater

clearly identifies their age. - Qualitative interpretation : measurable 3H= component of modern recharge - Radioactive decay : Calculating the time for decay from a known input level to measured

level - Exponential model for input function- compared to the previous approach, it assumes that

recharge is a composite of several years’ precipitation - Time series analysis-The bomb spike is identified through a sequential sampling of

groundwater for tritium analyse over several years. The three last approaches are based on the decay equation:

3Ht = 3Ht0 e-λ(t-t0)

where 3Ht and 3Ht0 are tritium activity at time t and t0, and λ is the 3H decay constant (equal to ln2 divided by the half-life t1/2 ). More than forty years after the thermonuclear bomb peak, the half- life of tritium implies that our ability to produce quantitative tritium ages for groundwater is really decreasing. • 3H/3He Measurement of 3H together with its daughter 3He, allows to enhance the use of 3H as a hydrologic tracer. With this method, true ages can be determined through calculation that do not rely on tritium input function. The 3H/3He age of a water sample is defined as:

t = λ-1 ln (3He*/3H +1)

where λ is the 3H decay constant, 3He* is the tritiogenic 3He concentration. It must be pointed out that 3He in groundwater has several sources: atmospheric helium, helium produced by nuclear reactions in the subsurface (derived from U and Th-series) and helium of mantle origin. The precision of the dating method depends on the accuracy of determination of the different He endmembers. A draw back of this method is that 3He is not a routinely measured isotope. • (CFC) Unwanted contaminant in our atmosphere, Chlorofluorocarbons (CFCs) are stable, synthetic halogenated alkanes developed in the early 1930s. Atmospheric CFC concentration have been increasing since the 1940‘s making them useful marker for modern groundwater. They are considered as excellent dating tools of young water at 50 year time scale.



Analysis of CFC compounds, made by gas chromatography and electron capture detection, is thus less complicated than for 3He. • 36Cl Thermonuclear bomb testing generated not only tritium, but also radioactive isotope of Cl- 36Cl, allowing its use to identify modern recharge. However, unlike tritium, its long half-life (301 000 ± 4000 years) precludes the use of 36Cl to date modern waters. • 85 Kr 85Kr has a half life of 10.76 years, thereby of potential interest for dating young groundwater. However, sampling (~100 l) and measurement preclude its routine use as an environmental isotope

b) Detecting submodern groundwater 3H/3He and CFC are useful dating tools for modern groundwaters. But, it is difficult to date submodern ones, in the range between 50 and ~1000 years BP. Unfortunately few radiogenic isotopes have a half-life in the range for these groundwaters. Argon-39 does (t1/2= 269 years), but its sampling and analysis is far from routine. Silica-32 also has an appropriated half-life time ((t1/2= 140±20 years), but is complicated by mixing with the abundant Si sources in the subsurface and because it does not behave as a conservative tracer.

c) Dating old groundwater For older groundwater, several tools are available. They include: - Stable isotopes of groundwater- Temperate latitude climates have experienced significant

change in temperature since late Pleistocene. Late Pleistocene paleogroundwater are isotopically depleted with respect to modern waters. This depletion allows to detect such groundwaters

- Radiocarbon- With a half-life of 5,730 years, 14C is a leading tools in estimating the age of paleo- and fossil groundwaters. The method is based upon the incorporation of atmospherically derived 14C from the decay of photosynthetically-fixed carbon in soils. Radiocarbon in the soil can be dissolved into water either as inorganic or organic carbon. Use of this tools implies to take into consideration the multiple sources and sinks of carbon in geosphere

- Chlorine 36- The long half-life of 36Cl (301,000 ± 4,000 years) and generally simple chemistry of Cl- makes this radioisotope an interesting tool for dating very old groundwaters, with application throughout the Quaternary period (1,650,000 years).

- Uranium decay Series – It includes: 234U/238U disequilibrium which allows dating of groundwater tens to hundred of thousands years- 226 Ra and its daughter 222Rn and accumulation of 4He

5.2.2.3. Residence time in unsaturated zone

As highlighted above, residence time of groundwater is highly variable from less than one year to several thousands of years. Depending on this residence time, transfer time of water and consequently of pollutants within the unsaturated (vadose) zone cannot be neglected.



High residence time illustrated by use of Tritium within the chalk Aquifer of Northern France (figure 5.3).

0

5

10

15

20

25

30

35

40

-60 -50 -40 -30 -20 -10 0 10 20 30 40 50 60

Dep

th (m

)

Tritium activities (TU)

Parcel A Parcel B

Figure 5.3 - Tritium profile within the unsaturated zone of Chalk Aquifer

(Hallue Basin - Northern France) (Normand et al., 1999). Such a figure suggests a high variability of transfer time - Over the two profiles, the 1963 tritium peak is recognized at two different depth suggesting

over parcel B, circulation is two times slower than on parcel A; - Patterns illustrates only slow circulation into the matrix porosity of the chalk and does not

preclude any rapid transfer through fissures. Figure 5.4 compares tritium and nitrates profiles in the moisture of a Loess (Baran et al., 2005) and illustrates how transfer of pollutants to groundwater may be delayed with respect to spreading time.



0 10 20 30 40 50 60 70

Tritium (TU)

15

14

13

12

11

10

9

8

7

6

5

4

3

2

1

0

Prof

onde

ur (m

)

0 40 80 120 160 200nitrate (mg/L)

TritiumNitrate

1963

1972

Figure 5.4 - Tritium and nitrate in soil moisture in a loess soil - France (Baran et al., 2005).

1963: year of maximum the tritium concentration in precipitation and 1972: year of commencing agricultural activities at the study site.

Low transfer time within the vadose zone, may have several implication regarding behaviour of pollutants, particularly the possibilities of reactivity, especially for oxidizable compounds. Finally it is worth pointing out that knowledge of transfer and reactivity of pollutants within the vadose zone needs additional research works. Concerning transfer time within the vadose zone, further research is needed to 1) have an overview of transfer time according to aquifer type and 2) to decipher between rapid and slow transfers. For example, in the case of chalk described above, the transfer time is probably the slow transfer within the matrix rock. More rapid transfer through fissures cannot be dismissed but we have no idea of the relative contribution of rapid and slow transfer to the recharge. The behaviour of pollutants in the vadose zone: negative pressure within the implies specific thermodynamic properties of water and dissolved compounds which are poorly known.

5.2.3. Flow processes – attenuation by dilution The main flow processes are listed in table 5.2.



Table 5.2 - Summary of important flow processes affecting solute fate and transport

(modified after Carey et al.). Process Description Dependencies Effect Recharge (simple dilution)

Movement of water across the water table into the saturated zone

Dependent on aquifer matrix properties, depth to groundwater, surface water interactions, and climate

Causes dilution of the contaminant plume and may replenish electron acceptor concentrations, especially dissolved oxygen

Advection Movement of solute by bulk groundwater movement

Dependent on aquifer properties, mainly hydraulic conductivity and effective porosity, and hydraulic gradient. Independent of contaminant properties

Main mechanism driving contaminant movement in the subsurface

Dispersion Fluid mixing due to ground water movement and aquifer heterogeneities

Dependent on aquifer properties and scale of observation. Independent of contaminant properties

Causes longitudinal, transverse and vertical spreading of the plume. Reduces solute concentration

Physical retardation

Trapping and retardation of solute due to mixing of water with immobile water zones

Dependent on aquifer properties and scale of observation. Independent of contaminant properties

Causes trapping of solute in immobile water zones followed either by other immobilization processes or by a delayed restitution

Partitioning from NAPL

Partitioning from NAPL into groundwater. NAPL plumes, whether mobile or residual, tend to act as a continuing source of groundwater contamination

Dependent on aquifer matrix and contaminant properties, as well as groundwater mass flux through or past NAPL plume

Dissolution of contaminants from NAPL represents the primary source of dissolved contamination in groundwater

Dilution may be an important process causing attenuation of pollutants. It can be considered at different scales and for different representative elementary volumes. Considering dilution at the whole aquifer scale shall require a good knowledge of recharge/discharge conditions while, at the opposite, a microscopic consideration of the dilution shall focus on the hydrodispersive characteristics of the aquifer. An example of macroscopic dilution can be the water entering the aquifer with a big discharge through sinkholes of karst areas or the dilution in an alluvial aquifer by the bank infiltration of the river. Dilution may also be closely linked to the part of the flow system considered (downward or upward flow). Advection is the process hinting any dilution, corresponding to the displacement “as it” (like in a piston flow movement) of the groundwater. It is sometimes predominant in karst conduits. Physical retardation may be observed in dual porosity media when the pollutant circulating through higher transmissive fissures or conduits may diffuse towards immobile water zones and be released with a delay. The mixing of waters with different geochemical contents may also have an influence on the physico-chemical attenuation. Advection may obviously have an impact on physico-chemical attenuation since it modulates the time of contact and the eventual exchanges (i.e. adsorption) between water and the aquifer. The hydrodynamic dispersion will increase the specific surface available for the interactions with the aquifer while, at the contrary, it will reduce these interactions by causing a lower concentration in water. Physical retardation is not often considered as far as it creates the same effect as sorption. However it is more linked to the drainage structure of the aquifer and may be a relevant



process for attenuation because it allows pollutants to spend a longer time in nearly immobile water with a high specific surface at disposal. The last process is a special case where an independent NAPL forms a third phase (solid, water, non aqueous-phase liquid) complexifying the flow.

5.2.4. Physico-bio-chemical attenuation in aquifers Physico-bio-chemical processes that are likely to attenuate pollutants have already been described in chapter 3, 5 and 6 of WP1 section and they will not be more detailed from a theoretical point of view. It has also already been mentioned that the effectiveness of these processes may vary according to the kinetics, the latest being itself dependent on the modalities of residence of the pollutants inside the aquifer. Many of the hydrogeological controls regulating flow will thus have an impact on these processes. The attenuation can be practically described using several distinct physical variables, like for instance: - the concentration C in groundwater; - the mass M of pollutants. While a decrease of both is clearly an objective, their dependance is sometimes not enough considered. The concentration may be evaluated at one single point in the groundwater body while the evaluation of the mass (or eventually the flux) requires the integration of the concentrations in a given volume. This simple statement is important to keep in mind when dealing with attenuation. For this reason it is common to distinguish degradative processes from non degradative ones. Both categories of these processes may lead to a reduction of the concentration along the pathway (flow line) but in case of degradative processes the pollutant is removed from the system (by a reduction of the effective mass) while in the case of a non degradative processes the pollutant is always subject to give a residual contamination by a delayed mobilization. The limit between degradative and non degradative processes is not always so clear when considering the reversibility or not of some reactions. Many of the processes involving mineral chemicals (majors, metals) are usually non degradative processes while many of the processes that concern organics are degradative. Physico-bio-chemical processes can be categorized as in table 5.3. Table 5.3 - Summary of important physico-bio-chemical processes affecting solute fate and transport

(modified after Carey et al.).

Process Description Dependencies Effect

Filtration Affects solid particles movement but may also become effective at a smaller scales

Depend on particle size and aquifer matrix properties

Immobilization of contaminants when they are in solid phase

Non

-deg

rada

tive

Sedimentation Affects solid particles movement

Depend on particle size and aquifer matrix properties

Immobilization of contaminants when they are in solid phase



Diffusion Spreading and dilution of contaminant due to molecular diffusion

Dependent on contaminant properties and concentration gradients. Described by Fick’s Laws

Diffusion of contaminant from areas of relatively high concentration to areas of relatively low concentration. Generally unimportant relative to dispersion at most groundwater flow velocities

Precipitation (dissolution)

Precipitation due to the limit of solubility of substances with which the pollutant ion can compose in groundwater

Dependent on contaminant properties and geochemical context.

Eventually causes the immobilization but may be reversible.

Sorption Reaction between aquifer matrix and solute whereby contaminants become sorbed or organic carbon or clay minerals

Dependent on aquifer matrix properties (organic carbon and clay mineral content, bulk density, specific surface area, and porosity) and contaminant properties (solubility, hydrophobicity, octanol-water partitioning coefficient)

Tends to reduce apparent solute transport velocity and remove solutes from the groundwater via sorption to the aquifer matrix

Cation exchange

Exchange of cations with elements present in the matrix of the aquifer, usually clays

Dependant on aquifer matrix properties and contaminant properties

Removes cations from groundwater. Can create a total removal or a simple delay if exchange is reversible

Volatilization Volatilization of contaminants dissolved in groundwater into the vapour phase (soil gas)

Dependent on the chemicals’ vapour pressure and Henry’s Law constant

Removes contaminants from groundwater and transfers them to soil gas

Abiotic degradation

Chemical transformations that degrade contaminants without microbial facilitation, such as hydrolysis, oxydation-reduction…

Dependent on contaminant properties and groundwater geochemistry

Can result in partial or complete degradation of contaminants. Rates typically much slower than for biodegradation

Deg

rada

tive

proc

esse

s

Biodegradation Microbially mediated oxidation-reduction reactions that degrade contaminants

Dependent on groundwater geochemistry, microbial population and contaminant properties. Biodegradation can occur under aerobic and/or anaerobic conditions

May ultimately result in complete degradation of contaminants. Typically the most important process acting to truly reduce contaminant mass

Most of these processes are detailed from an hydrogeological point of view in Kehew (2001), Fetter (1993) or Carey et al. (2000). Filtration and sedimentation are more physical processes. They have been included here because they involve a solid phase which can hardly be classified in the pure flow processes. They require firstly the fixation of the pollutants onto a solid phase through any of the physico-bio-chemical processes (precipitation, sorption…). Obviously the transport of particles also play an important role in the attenuation of numerous sorbed pollutants (metals, organics, micro-organisms). Sorption covers a wide range of situations that may be more or less reversible. The reversibility of any process depends largely on the possible changes that may occur in the geochemical context of the aquifer or along the pathway. Here again hydrogeological controls are crucial. Cases where one pollutant is transformed into another more or less harmful substance are also common, especially for organics.



5.2.5. Hydrogeochemical background The hydrogeochemical background has already been presented in the chapter 4 of WP1 section and will be more detailed as well in the chapter 6 of the present section. The background may also be used for assessing the geochemical conditions controlling the attenuation of any pollutant, not only naturally occuring substances. For this reason, in the present chapter, the background will be considered as a controlling factor among others. This factor can obviously be related to the type of aquifer. The salinisation of an aquifer may also be considered as inherited or more recently produced, depending on the time scale considered. It is not rare that deep aquifers contain old aged waters issued from old marine waters, even diagenesis waters. 5.3. Aquifer typology and attenuation of pollutants

5.3.1. Principle and objectives of the typology Each type of aquifer presents specific characteristics conditioning the fate and particularly the attenuation of pollutants. Consequently, the typology defined within chapter 3 will serve as a basis in this section to describe the main characteristics of aquifer which may control the fate of pollutant. A pragmatic typology could thus be a simplification of the -too high- diversity of aquifers which may help MS to determine threshold values in a pretty harmonized way. There is indeed a so wide variety of aquifers and they are controlled in a so wide manner that it seems difficult to establish now a strict typology. The general typologu applied here is only a first approach. It is thus left to the appreciation of the MS to use their own conceptual model -as mentioned in the annex III of the Common position adopted with a view to the adoption of the groundwater directive- to better derive the potential attenuation of the pollutants in their specific hydrogeological situations. The typology must reflect as far as possible the nature of flow involved in the aquifers. The combined lithological/stratigraphical features may often reflect this flow typology because they are related to the primary and secondary porosity which determine the organization of the drainage inside the aquifer. It must also reflect the potentiality of physico-bio-chemical processes according to the geochemical content. Note that a simplified typology can be inversely used in a first step for approaching aquifers where characterisation data are lacking. Data on the attenuation of pollutants in different aquifers were collected among the partners, trying to cover an as wide as possible range of aquifers. The required answers included the following information: - general information on the location and dimension of the aquifer at a European scale; - geological settings (lithology, stratigraphy…); - hydrogeological (physical) properties; - cover; - flow conditions;



- physico-bio-chemical conditions; - description of all physico-bio-chemical processes (including their controls) identified

according to any relevant pollutant. The table 5.4 summarizes the main characteristics of the described aquifers. Besides the typical characteristics, are also mentioned the pollutants (whether from natural or from anthropogenic origin) that have been characterised from a processes point of view. A synthesis on each type of aquifer is presented in the following chapters. Each section is divided into 2 subsections. The first one focus on flow characteristics of the aquifer type whereas the second one deals with the geochemical characteristics. Attenuation conditions are fully dependant on the type of pollutants - for details on a specific compounds, reader is invited to consult WP1 report. The second point describes the conditions which may be of relevance regarding the fate of some pollutants, namely: (1) redox condition as some pollutant may degrade under oxidizing conditions whereas other may degrade under reducing condition; (2) adsorption capacities of the aquifer as some polluants may adsorb on various phases (e.g. oxides, clays) (3) the pH of the groundwater as solubility of some compounds depends on pH; (4) occurrence of carbonates as some element may precipitate or co-precipitate with carbonates (5) complexation capacities which may enhance solubility of some minerals (e.g. presence of Cl, SO4 in groundwater).

5.3.2. Sandstones - siltstones aquifers

5.3.2.1. Flow characteristics Intergranular porosity of sandstones aquifers is generally quite low (<10%). Thus, most of the porosity in these consolidated rocks consists of secondary openings such as joints, fractures and bedding planes. According to their tectonic history, aquifers are more or less affected by this second porosity. As mentioned for the Triassic sandstones-siltstones of Germany, groundwater flow is primarily along bedding planes but joints and fractures allow a vertical movement to the water and provide avenues in the sandstones. This may play an important role on the attenuation of pollutants. In aquifers described by the partners, total porosity ranges from less than 5% to 35% in most affected zones, the effective porosity being lower. Sandstones generally consist of layered rocks differentiated by grain size, alternating low and high permeability layers. This effect appears in Germany (Triassic sandstones and siltstones), United Kingdom (Permo-Triassic sandstones) and Portugal (Meso-Cenozoic sandstones) and usually results in a high vertical anisotropy. The hydraulic conductivity mentioned by partners ranges from 1.2.10-11 to 2.3.10-4 m/s while residence times take values between 10 years and several hundred years. In Estonia (Cambrian-Vendian aquifer system), groundwater recharged during the last glaciation has been preserved in the aquifer.



Tabl

e 5.

4 - S

umm

ary

of m

ain

char

acte

ristic

s of

the

Eur

opea

n aq

uife

r and

rele

vant

sub

stan

ces

for w

hich

pro

cess

es w

ere

best

iden

tifie

d.



5.3.2.2. Geochemical characteristics

In some aquifers, conditions are dominantly oxidizing with abundant iron oxides. These oxic facies include: Permo-Triassic sandstone (UK) and Red-Bed Devonian sandstone (part of Wales and welsh Borders) and Buntsandstein sandstone from Germany. In other places, reducing conditions prevail- they are mainly illustrated through NO3 disappearance. Within the Aveiro Cretaceous aquifer, NO3 depleted water are encountered as soon as aquifer is confined and redox processes are considered as important in the confined part of the Permo-Triassic sandstone aquifer of UK and in the Carboniferous Millstone Grit groundwaters (UK). On one hand, to explain these redox conditions pyrite is reported within Greensand deposits and poorly- consolidated tertiary sand aquifers (UK), and on the other hand organic matter, but at very low level, is recognized within the Permo-Triassic sandstones aquifer of UK. Regarding adsorption capacities oxides phases were recognized under oxidizing conditions and clays, may be present. For example in Greensand deposits in southern England. among others glauconite has been recognized. Within the Aveiro Cretaceous aquifer, cation exchange is considered as a main process responsible for chemical evolution. Carbonates minerals are also encountered in such aquifer – almost all of them contain carbonate minerals, but in varying proportion. Dolomite is present at small percentage (typically 1-4wt%) within Permo-Triassic sandstone. Carbonates content of Greensand deposits of southern England is typically around 1wt% and variable amount have been reported in Tertiary formations from southern England. Within this last formation, carbonates minerals are suspected to control concentrations of Pb and Cd. Mineral dissolution (Gypsum, anhydrite) and/or oxidation (pyrite) as well as saline intrusion may increase SO4 concentration has been observed in UK, enhancing potential complexation of some pollutants. Within the Aveiron cretaceous aquifer, pH is highly variable, with values from 5 to 9.15. On the other side, groundwater from Triassic sandstone of Germany are threatened by acidification due to thin and buffer poor soils. This more than 50% of all samples display a pH less than 7.0.

5.3.3. Sands and gravels aquifers Referring to the classification by aquifer type, “sands and gravels” consist in a class gathering pure sands (e.g. Neogene and Paleogene aquifer of Flanders), pure gravels (e.g. fluviatile deposits in Germany), sands and gravels (e.g. glacial deposits in Germany) and multilayered aquifer system (e.g. quaternary deposits of Hungary, Spain and Portugal). In a geochemical point of view, this choice is optimal, referring mainly to the chemical composition of the aquifer. However, in a hydrogeological point of view, great differences are likely to be observed, according to aquifer characteristics. On several occasions, a subdivision of this class will be necessary.

5.3.3.1. Flow characteristics Hydraulic conductivity of an aquifer strongly depends on granulometry and porosity. In examples treated by partners, it varies between 1.10-5 and 6.10-2m/s. In the case of



multilayered aquifer systems, hydraulic conductivity is very heterogenic and can be lower than 1.10-7m/s in clayey layers (intergranular multilayered aquifer system of Hungary). In the same way, total porosity takes value between 20 and 50% while effective porosity varies between 5 and 30%. As shown in the Hungarian contribution, total porosity of a clayey layer can reach high values but the effective porosity (which is important for the waterflow) is generally very low (respectively 60% and <10% in the Hungarian intergranular multilayer aquifer system). These clayey layers can thus provide a good protection for the subjacent aquifer. The nature of the cover of the aquifer is extremely variable of an aquifer to the other and even above a single one. In Portugal (multilayer system), a very thin layer does not make it possible to protect the aquifer whereas the clayey cover of the fluviatile deposits aquifer in Germany provide a relatively good protection. Due to low velocities, residence time in porous aquifer can reach great values. It also depends on the flow system scale. Thus, shorts residence times are low for local flow system (quaternary deposits of Germany) and quite high for regional flow system (Paleogen and Neogen sands of Flanders, Multilayered system of Hungary and Portugal). It varies between 1 year and several hundreds years.

5.3.3.2. Geochemical characteristics Both oxidizing and reducing conditions are reported within this type of aquifer. Under oxidizing conditions, pollutant occurrences include NO3, but also SO4, K and organics compounds in places. It is observed whatever the aquifers: Aveiro Quaternary Aquifer (Holocene and Plio-Pleistocene deposits from Portugal), Sands-Neogene aquifer (Flanders), Ledo-Paniselian aquifer (Flanders), Intergranular multilayered aquifer system from Hungary, Plana de Castellon aquifer (Spain) or Glacial and sand gravel deposits from north Germany. Such oxidizing conditions prevail within the Aveiro Quaternary aquifer (Portugal). However in various conditions (confined, but also shallow unconfined conditions), redox reactions can be active. Denitrification is then generally reported, which illustrates the capability of reduction within the aquifer. In some cases, stronger redox conditions can be also reached. For example, in the multilayered aquifer system from Hungary, redox sequence goes up to sulphate reduction and even to methane fermentation in Sands-Neogene aquifer (Flanders). Both pyrite and organic matter in the solid phase can act as electron donors for these reactions. Presence of pyrite is mentioned for Sands-Neogene Aquifer (Flanders), Ledo-Paniselian aquifer (Flanders), intergranular, multilayered aquifer system from Hungary, or Plana de Castellon aquifer (Spain). Organic matter has been recognized in Sands-Neogene Aquifer (Flanders), intergranular, multilayered aquifer system from Hungary. Moreover, the rather high DOC values in the groundwater of Aveiro Quaternary Aquifer (Holocene and Plio-Pleistocene deposits from Portugal) allow envisaging organic matter occurrence as solid phase. Other minerals like glauconite and siderite through dissolution of Fe(II) may contribute to the reduction of species (illustrated for Sands-Neogene Aquifer (Flanders)). Transfer of electron donors through water can contribute through reduction of pollutants. In the Pannonian Basin, organic matter is dissolved during circulation of water at depth. However, in the discharge areas, the organic carbon bearing upflowing waters mix with shallow groundwater contaminated by agricultural activities. The transported organic carbon is available for reduction of pollutants as illustrated for nitrates.



Consequently to reducing conditions, occurrences of NH4 and NO2 have been reported, NH4 is due to organic matter decomposition in the groundwater of Intergranular, multilayered aquifer system from Hungary. Both elements may be a consequence of NO3 reduction (Intergranular, multilayered aquifer system from Hungary, Ledo-Paniselian aquifer (Flanders), Sands-Neogene aquifer (Flanders)). It is worth pointing out that despite the geogenic origin of some elements, high concentration may reflect anthropogenic impact. For example Arsenic, dissolved from sulphide minerals, has been reported in Intergranular, multilayered aquifer system from Hungary and, Sands-Neogene Aquifer (Flanders). Oxidation of sulphide minerals by NO3 reduction may contribute to such an increase. When iron of vivianite (Fe3(PO4)2.8H2O), presents for example within Sands-Neogene Aquifer (Flanders), is involved for denitrification or any other reduction, PO4 concentration can increase rapidly. Reported pH values are highly variable: values from 3.5 to 9 are reported for the Sands-Neogene aquifer (Flanders), and between 4.6 and 7.6 within the groundwater of Aveiro Quaternary aquifer. In such aquifer type, carbonate phase content is highly variable: it can be totally absent or present in significant amounts. For example, calcite is present only in a part of the Neogene aquifer (Flanders).

Figure 5.5 - Structure and geochemical characteristics of sands- Neogene aquifer

(Northeast Flanders). Sorption/desorption, ion exchange capabilities are due to presence of clay minerals. But, in oxidizing conditions, occurrences of oxide-hydroxide phases such as iron hydroxide and Mn-oxides (Ledo-Paniselian aquifer, Flanders) have been reported. Within the Intergranular, multilayed aquifer system of Sofia valley, adsorption of Pb, Zn and As on clays is reported Saline intrusion / component are reported in some of these aquifers and illustrate capabilities of the aquifer. The Ledo-Paniselian aquifer consists in initial marine conditions and is freshened by precipitation recharge. This freshening involves an increase of marine cations



(Na, K, Mg as well as NH4 formed at the bottom of the sea by reactions with organic matter) and a decrease of Ca. Cation exchange has also been reported within the intergranular, multilayered aquifer system from Hungary. High sulphate concentrations are reported within groundwater of the Intergranular, multilayed aquifer system of Sofia valley, which may have an impact on solubility of some metals as mentioned for iron.

5.3.4. Limestones (karstic/non karstic) aquifers Chalks owing a considerable "additional" intersticial porosity compared to other carbonate rocks, they are dealed with in a separate section. In the present section only limestones formations, including interbedded dolomites, sandstones and shales, are described.

5.3.4.1. Flow characteristics The karstification alteration may affect all kinds of carbonate rocks, including chalks, and in some specific cases also non carbonate rocks (evaporites, sandstones, metamorphic rocks…). Karstic features are thus clearly an important criterium that must be used in a second step in the typology among carbonate rocks. The specificities of karstic aquifers are described in an abundant literature, among others in Worthington (2003), Kiraly (2003) and Bakalowicz (1999). The distinction between karstic/non karstic is a strong debate among karst scientists. Indeed there is a continuum between diffuse, fractures and conduit flow patterns (see figure 5.3, after Quinlan & Ewers). The main distinction often used is that the Darcy's law may be applicable to non karstic rocks while it is no more applicable for karstic conduits. The behaviour of springs or cave streams is also often used as a distinction, being "flashy" (Quinlan, 1989) in the case of karstic flow and more "inertial" in the case of non karstic functioning. Also the turbidity may be used as an indicator of the karstic behaviour12. The karstic behaviour is even more complex since karstic voids may exist without necessarily being functional. Thus, in some cases, the karstic behaviour may change suddenly according to flow conditions (whether they are anthropogenic or natural). The variability in both volume and velocity of water through conduit systems may have a considerable impact on contaminant transport (Vesper et al., 2001). Karst aquifers have also some peculiarities like the existence of a groundwater buffer zone close to the surface which is called epikarst and that may play an important role in the storage and transfer of contaminants. The thickness of the cover may therefore vary considerably, what is an actual challenge for estimating the vulnerability and the risks for such aquifers. The table 5.4 gives the main characteristics of the aquifers described by the partners. Non karstic aquifers (Spain, Portugal, Germany) belong exclusively to mesozoic sequences. Due to their geological history (sedimentation as well as late tectonics) these series were not subject to the karstification. The karst genesis may be very complex but it is 12 Note here that the turbidity can also control the transport of many contaminants.



often the conjunction of favourable chemical and hydraulical factors that fix the development of the karst. These factors were not achieved in these three cases. The whole aquifer reservoir usually hold both the transmissive and storage functions unto very small scale. Velocities are usually low and the residence time long (tenths to hundreds of years). Even if these aquifers can include some enlargement of the fissures, that confers a multiple porosity, they can often be considered as homogeneous from a flow point of view. At the microscopic scale, the exchanges of solutes (or non aqueous phase liquids) between the fissures and the matrix may play an important role on the attenuation of the pollutants. The discharge of these aquifers is usually through linear springs, thus eventually impacting a great deal of a river. Karstic aquifers (Hungary, Germany, United Kingdom and Belgium) belong either to mesozoic or to paleozoic sequences. Those sequences have a higher degree of karstification due either to a higher content in carbonate (less interbedded shales or sandstones) or a higher degree of fracturation, sometimes due to successive orogenesis (hercynian plus alpine). For mesozoic sequences, on the contrary of non karstic mesozoic sequences, huge mesozoic sequences with less interbedded terrigeneous rocks, with a stronger fracturation or/and with a higher potential of karstification, were intensively karstified along the alpine belt13. In these aquifers, the transmissive and the storage functions are usually well dissociated and the drainage is mainly discrete through karst conduits where attenuation processes may be radically different from that of the remaining part of the aquifer where only a diffuse flow is present. The dilution along the conduits can be taken into account just as for the dilution in surface waters. As previously said the most important factor to consider is the variability of the processes. Velocities along the conduits may reach several hundreds of meters per hour. As the consequence of the strong drainage, residence times may range between several days for water entering the system through swallow or sinkholes to several years for precipitation water infiltrating. A mean residence time is difficult to define due to the superposition of several flow components. As a consequence, sources of pollution in karstic aquifers may often be considered exclusively as point source pollution. Entries of contaminants may be sudden and intense and often coming from surface streams, what is a reason for monitoring surface waters in order to protect karst water resources. Karstification may also be due or facilitate the entry of thermal waters into the carbonate aquifer like in some part of the limestones aquifer of the Transdanubian Central Mountains in Hungary. It is observed, like in Wallonia, that a same aquifer (indeed carboniferous limestones) may behave like a less or a more karstic one according to its depth. Deep confined parts of that aquifer seem to be less subject to an active karstic drainage than the same rocks exposed to a surface drainage system. The wavelength and amplitude of folding or faulting are also factors for the development of a karstic drainage. 13 The example of limestones in South of France is not described here though it is rather typical.



5.3.4.2. Geochemical characteristics Both oxidizing and reducing conditions are also encountered within this type of aquifer., once again Organic matter and sulphide minerals are potential electron donors for reducing conditions setting up. Organic matter is present within the Jurassic Limestones, Magnesian Limestones and Carboniferous limestones of England and Wales, as well as in the Carboniferous limestones in Belgium, where it induces a denitrification of groundwaters especially in confined areas. On the opposite, low organic content is reported in the karstic aquifer of the Transdanubian Central Mountains in Hungary, and oxic conditions prevail in karstic carbonate aquifers of Germany. Sulphide minerals, among others pyrite, have also been recognized in some of these aquifers: Oxfordian Limestone from France, where pyrite allow denitrification within the confined part of the aquifer, and the Carboniferous limestones in the Tournaisis (Belgium). In such aquifers, SO4 concentrations may increase due to nitrates reduction. SO4 concentrations Increase is then likely to enhance solubility of some metals. Clays minerals are also reported in variable amount in such type of aquifer: Jurassic Limestones, Magnesian Limestones, Carboniferous Limestones of England, Wales and Wallonia may contain high clay contents often present as interbeds whereas they are rare in the karstic aquifer of the Transdanubian Central Mountains in Hungary. Compared to other types of aquifer, carbonate aquifers are characterized by higher pH values. The impact of these high values may both concern the equilibrium with respect to carbonates and the rate of redox reactions. For each pollutant, the rate of biodegradation depends on the pH value14.

5.3.5. Chalk aquifers

5.3.5.1. Flow characteristics As mentioned above, chalks are specific carbonate rocks (even limestones) with an enhanced diffuse flow capacity due to fine-grained texture (the total porosity can reach more than 50%). Chalks aquifers provide important groundwater resources in UK, north of France and south of Belgium. Chalks are usually gently dipping beds that may reach a thickness of several tenths of meters. It is characterised by a double porosity (fractures and matrix). Water velocity remains very slow in the matrix but a very efficient drainage of this matrix is allowed through solution-enlarged fissures of varying apertures. Though these apertures are never so wide and the situation never so contrasted than in other limestones, chalks may also be subject to a karstification. This karstification may be observed especially along the coasts or the main valleys (Rodet, 2004). It is not so hierarchised than in limestones, what is mainly due to a lack of coherence of this type of rock.

14 For information, the reader is invited to consult data acquired in the frame of WP1.



However, it is typical to find greater fluxes of water along dry valleys, where an deeper alteration could take place. Thanks to a common higher "radius" of influence of these heterogeneities, they are often taken into account in groundwater models (Hallet, 1998) in a better way than it could be done for classical karst. Indeed, lower water velocities allow the Darcy's law to be used. Chalks may contain flint, marls horizons or hardground that may play an important role in the organization of flow. Chalks are mainly exploited through water wells with a considerable yield or through extended linear galeries (Hallet, 1998). The recharge is essentially a diffuse one but swallow holes (called "bétoires" in France) may also exist. The vertical infiltration is dominated by a diffuse flow that allows many geochemical process to happen (great exchange surface and residence time). Chalks are far more vulnerable when they are unconfined, shallow aquifers than in confined position.

5.3.5.2. Geochemical characteristics The Cretaceous Chalk in UK, France and Belgium presents low organic matter content, in the range of <0.01-0.1wt%. This content may however increase in the marly beds. Nevertheless, in the confined parts, conditions are reducing and denitrification has been observed. In the Chalk aquifer of Northern France, denitrification has also been evidenced in the confined parts of the aquifer: both pyrite and organic matter are supposed to contribute to this process. Sorption/desorption processes are suggested to explain Pb and Cd control within groundwater.

5.3.6. Schists and Shales aquifers The term “shale” designates any compacted clays presenting a fissility (plitting easily in thin folders) parallel to the stratification. Usually, the term schist is wrongly employed while it designates an indured fine-grained rock affected by a schistosity (cleavage due to the dissolution and the simple reorientation of minerals under tectonic pressures). For this type of rock, refers to “crystalline aquifer” section.

5.3.6.1. Flow characteristics Shales aquifer has not been treated by the partners.

5.3.6.2. Geochemical characteristics For schists aquifer, the reader is requested to refer to “crystalline aquifer section” below . Compared to crystalline aquifers, occurrence of sulphide minerals and organic matter may be different.



5.3.7. Crystalline aquifers This category of rocks includes magmatic and metamorphic rocks.

5.3.7.1. Flow characteristics Those aquifers are characterised by a strong contrast between the permeability of the matrix and the permeability of the fractures (fissures). Magmatic and metamorphic rocks are particularly hard rocks that consequently have a specific, discrete tectonics that creates a typical secondary openings network. Unlike karst aquifers, this network can hardly be enlarged by any solution process. The connections between the different voids are rather poor and usually remain as it. This means that the drainage can not be progressively organized unto large springs like it is the case in the karst, where the evolution may be very rapid. Hydrogeological characteristics (e.g. hydraulic conductivity and storage) derive primarily from the geomorphologic processes of deep weathering. Weathering processes occurred mainly during Early Cretaceous and Tertiary. The classical profile is composed of the following layers which have specific hydrodynamic properties (Wyns et al., 2004; Maréchal et al., 2005):

1) Unconsolidated weathered rock (saprolite or regolithe), derived from prolonged in-situ decomposition of bedrock, (High porosity depending of the clay content and low permeability zone);

2) Fissured- weathered layer, generally characterised by a fissure density that decreases with depth. This layer insures the transmissive function of the aquifer;

3) Fresh basement, which is permeable only where deep tectonic fractures area present. Permeability varies with lithology, Rough estimates fomr France and germany are presented below: - granites and gneiss: 9,0.10-6 to 1,0.10-7 m/s; - sandstone - quartzites - sandstone and schists: 1.10-5 to 9,0.10-9 m/s; - schist (37): 6.10-5 to 1.10-8 m/s. The residence time of water may vary with depth and distance from the fissures, from typically a few years to more than 50 years.

5.3.7.2. Geochemical characteristics At shallow depth (saprolite), oxidizing condition prevail. Redox conditions in these aquifers depend mainly on the lithology. In plutonic igneous rock, like granite, sulphide minerals or organic matter are rare and oxic conditions prevail. The attenuation of pollutants like NO3 is very low, if not absent, although denitrification through Fe(II) of unidentified minerals have been reported from laboratory experiments using granite ships. However, in the structural contact zones, accumulations of pyrite have been recognized and very efficient denitrification process has been evidenced.



In metamorphic rocks, like micaschists, gneiss or schists, sulphide minerals and even organic matter are more common. In the Brioverian schists of Brittany (France), only pyrite has been recognized. It allows denitrification at a very rapid rate: the half life of nitrate is in the range of 2.9 to 7.1 days, according to the medium of circulation of contaminated water (porous matrix of the aquifer or fissures). Low pH values of these waters as reported for groundwaters of crystalline aquifers in Portugal, Germany or France. This favours the solubility of some metals. As long as redox conditions allows iron oxidation, adsorption onto oxides phases can control concentrations of some (but not all) metals/or metalloids. In Brittany (France), subsequently to denitrification, induced by sulphides minerals, increases of As, Zn, Mo, Pb, Cu concentrations are expected.

5.3.8. Volcanic rocks

5.3.8.1. Hydrogeological characteristics In the Colli Albani structure (Italy), water circulation develops in radial direction from the centre to the periphery following complex patterns. The geological settings originated an aquifer in the central area sustained by low hydraulic conductivity rocks and a basal aquifer sustained by marine pre-volcanic clay deposits and contained in more ancient volcanic rocks. Water also circulates through lakes from the superior to the basal aquifer. In Germany, most productive aquifers consist of basalts originating from the tertiary. Basaltic lavas tend to be fluid and they form thin flows that have considerable pore space at the top and the bottom of the flows. Different flows are separated by paleosols or alluvial material that forms permeable zones. Columnar joints allow water to move vertically through the basalt. In Hesse (Germany), Hydraulic conductivity of volcanic rocks ranges from 10-6 to 10-5m/s. Residence time of water takes value between 5 years to several hundreds years in deep lying aquifer.

5.3.8.2. Geochemical characteristics In Germany, waters in volcanic areas are mainly oxic.whereas the Italian Colli Albani structure aquifer is concerned by rising gasses rich in CO2, SO2 and H2S. Through their oxidation, SO2 and H2S gases can contribute to reduction of species, among them potential pollutants according to their chemical properties. Within the Massif Central of France, where basalt outcrops, low ranges of denitrification of water have also been evidenced, although in this case, origin of the process has not been documented. In Italy, oxidation induces also a strong acidification of waters, which pH around 3, which greatly modify the solubility of species like Fe, Al and As. At this stage, origin of As is not recognised as natural or not. However, effect of pH on potential pollutants solubility cannot be dismissed. On the opposite, pH of water in Germany are near neutral to basic. In Italy, gases oxidation also induces increase of SO4 concentrations which in turn may impact solubility of some metals. In volcanic area of Germany, occurrence of zeolite leads to high ionexchange capacity.



5.4. References Baran N., Bourgeois M., Flehoc C., Normand B. (2005). Détermination de la vitesse de transfert de l’eau, des nitrates et autres solutés en zone non saturée dans un lœss profond. BRGM/RP-53440-FR, 88p, 46 ill. Bakalowicz, M. (1999). Connaissance et gestion des ressources en eaux souterraines dans les régions karstiques. Guide Technique n° 3. Bassin Rhône-Méditerranée-Corse. Bear, J., Verruyt, A. (1987). Modeling groundwater flow and pollution. Theory and applications of transport in porous media. Reidel, Dordrecht. 414 p. Carey, M.A., Finnamore, J.R., Morrey, M.J., Marsland, P.A. (2000). Guidance on the assessment and monitoring of natural attenuation of contaminants in groundwater. U.K. Environment Agency. R&D Publication 95. 131 p. Fetter, C.W. (1993). Contaminant Hydrogeology. Mac Millan Publishing Company. New York. 458 p. Cost 620 Cost Action 620, final report (2003). Vulnerability and risk mapping for the protection of carbonate (karst) aquifers. Zwahlen Ed. 297 p. De Marsily, G. (1986). Quantitative Hydrogeology. Academic Press,INC- 440p. IAEA (1990). Guidebook on radioisotope tracers in industry, Technical reports series n° 316, Vienna (1990), 374 p. Kehew, A.E. (2001). Applied Chemical Hydrogeology. Prentice Hall, 368 p. Kiraly, L. (2003). Karstification and groundwater flow. Speleogenesis and evolution of karst aquifers, 1 (3), September 2003, 1-26. Maréchal J.C., B. Dewandel, K. Subrahmanyam (2005). Hydrodynamic characterization of the weathered-fractured layer of a hard-rock aquifer: (1) geometry and permeability of fractures network Normand B., Czernichowski-Lauriol I., Mouvet C. (1999). Programme régional expérimental de suivi de la qualité des eaux sur trois bassins versants de Picardie faisant l'objet de mesures agri-environnementales réduction d'intrants. Suivi de la nappe et de la zone non saturée dans le basin de l'Hallue (Somme). Rap. BRGM R 40616, 321 p., 112 fig., 39 tabl., 3 ann., 3 c. Quinlan, J.F. (1989): Ground-water monitoring in karst terrains: recommended protocols and implicit assumptions. U.S. Environment Protection Agency, EPA 600/X-89/050, 79 p. Rodet, J. (2004). Karst et Craie en Normandie: une approche géographique. Karst and chalk in Normandy: a geographical approach. In: Le karst de la craie en Normandie. Actes des Journées Européennes de l'AFK, Rouen, 2003, 17-32. Vesper, D.J., Loop, C.M., White W.B. (2003). Contaminant transport in karst aquifers. Speleogenesis and evolution of karst aquifers, 1 (1), January 2003, 1-8.



Worthington, S.R.H. (2003). A comprehensive strategy for understanding flow in carbonate aquifer. Speleogenesis and evolution of karst aquifers, 1 (2), April 2003, 1-11. Wyns, R., Baltassat, J.M., Lachassagne, P., Legchenko, A., Vairon J. and Mathieu F. (2004). Application of SNMR soundings for groundwater reserves mapping in weathered basement rocks (Brittany, France). Bulletin de la Société Géologique de France 175(1) (2004). 21-34.



Chapter 6

Natural background levels. State of the art and review of existing methodologies

A. Blum (BRGM)

F. Wendland, R. Kunkel (fz-Jülich) Marleen Coetsiers, Marc Van Camp, Kristine Walraevens (LAGH-UGent)

Rossita B. Gorova (EEA) G. Berthold, G. Fritsche (HLUG)

Rüdiger Wolter (UBA-D) Klaus Hinsby (GEUS) Andres Marandi (UT)

Zoltan Simonffy (BME) Kestutis Kadunas, Jurga Arustiene (LGT)

Jasper Griffioen (TNO) Stanislaw Witczak (DHWP/AGH)

Jan Hookey (EA) Juhani Gustafsson (Finnish Environment Institute)



6.1. Introduction One of the main objectives of the Water Framework Directive (WFD) is to prevent and limit pollution. But how to identify and characterise pollution without having define before the natural background levels of the groundwater bodies (GWB)? In this context, the identification of the natural background levels will be an initial step for the definition of the good chemical status required by the WFD. Besides, it is required for the further characterisation process (article 5, annexe II.2). As agreed in within the BRIDGE project, only the term “natural background level” (NBL) will be used instead of “baseline concentration” or “background concentration”. The following definition has been accepted by all: “The concentration of a given element, species or chemical substance present in solution which is derived by natural processes from geological, biological or atmospheric sources”. Although the methodology helping to define the good chemical status is developed in WP3, this chapter aims to give a support for the identification of natural background levels which is a part of this methodology. As a complement of WP1’s report dealing with the chemical properties of each substance, this chapter aims to give a description of the “typical natural background levels” of each type of aquifer. Thus, for each type of aquifer, a description of this “typical chemical composition” is proposed, first for major elements, and then for trace elements. But before describing these results, it appeared important to deal with the different existing ways of assessing NBL through Europe. Many studies have indeed already been performed in many countries and the methodology developed within the BRIDGE project will need to take this knowledge into account. Furthermore, the level of knowledge is very different from a country to another and this chapter also aims to describe this heterogeneity that should also be considered in WP3 when deriving the methodology.

Figure 6.1 - Contributing countries.

Remark about the source of information: most of the information presented in this chapter comes from the contributions sent by WP2’s partners. A questionnaire circulated and thanks to the contributions of 11 countries (figure 6.1), a lot of information was available. As for



practical reasons it appeared impossible to include the all of each contribution, only the key points are reported in the core text. For more detailed information about each of these 11 countries, please refer to the appendix where the entire contributions have been included. 6.2. Approaches to determine Natural Background Levels Notice: this section aims to give an overview on how NBLs are approached and assessed in European countries. It clearly doesn’t aim to judge if a methodology is better than another one. The main goal is to show what has already been done and to underline the different existing approaches and the different levels of knowledge.

6.2.1. Main approaches Natural background levels are very differently considered through Europe. Some countries have developed national approaches (FR, D, BU) while others focused on regional scale studies (UK, DK, BL-Flanders, EE, NL, HU, PT). There are also countries in which, even if no shared national methodology has been defined, many works on the subject has been realized and guidelines helping to identify NBL are available. Estonia, Nederland, Poland Lithuania and Flanders thus defined limit values or background standards for trace elements.

Figure 6.2 - The different approaches to assess NBL through Europe. These two ways of defining NBLs are summarised in the following section. In appendix of the document, the contributions sent by each country are presented. More details on each country are thus available.

Local scale approach (aquifer)

National scale approach (typology)



6.2.2. National approaches (FR, D, BL, LT)

6.2.2.1. General concepts France and Germany have a similar approach. In both countries, a national scale statistical analysis has been performed in order to define the typical chemical composition of each type of aquifer. In Germany, neither than in France, “natural” groundwater is practically not present anymore. Human pressures are so important that it has been impossible to identify some aquifers that could have been selected as reference systems. In these conditions, Germany proposes to evaluate a number of groundwater samples from different locations within a homogeneous hydrogeological unit. In Bulgaria, two different studies have been performed to assess the chemical composition of fresh groundwater. Both meanwhile adopts a statistical approach and, even if no one is presented as a national and adopted approach, they are quite similar to the German and the French approaches and will be presented here. The main difference between these two methodologies is that one includes samples from mining area. In Lithuania, although no national approved methodology exists, a statistical approach using data available for the whole country has been performed. Common principles have been derived to assess NBL.

6.2.2.2. Summary of each approach • Germany In order to identify NBL and more specifically to distinguish natural levels from concentrations influenced by human activities, Germany developed a statistical approach based on the study of the concentrations distribution. This approach has been applied to a selection of data (see following paragraph) and NBLs have been derived for each type of aquifer. The main principles and concepts of this approach are explained here. Because of the natural variability, concentration measurements of a certain groundwater parameter from a number of sampling points within a homogeneous groundwater unit lead to a concentration distribution. In case no anthropogenic intakes to the groundwater are present, the shape of this distribution is determined exclusively by the natural variability of reactive exchange of groundwater with the aquiferous rocks. Typically, right-skewed asymmetric concentration distributions were observed in these cases. Several statistical tests performed in this study showed that the shape of the concentration profiles can be described very well by lognormal distributions (Kunkel et al., 2004). If the whole aquifer is ubiquitary affected by anthropogenic intakes the concentration profile of a certain parameter is modified. Because the concentration rise of the anthropogenic intakes is independent of the natural solution content of the groundwater, the concentration distribution of the ubiquitary affected aquifer can be regarded as a result of the convolution of the natural content distribution fnat and the distribution of the anthropogenic intakes fanth:

( ) ( ) ( )∫∞

ξ⋅ξ−=0

anthnatlinf dcfcfcf (1)

This is schematically sketched in figure 6.3. With respect to the individual contributing concentration distributions the convoluted influenced distribution is broadened and shifted toward higher concentrations. As a consequence of the convolution, all information about the



individual contributing concentration distributions are lost. Therefore, it is impossible to judge whether an observed distribution function of a certain groundwater parameter is still natural or ubiquitary influenced by anthropogenous intakes.

Figure 6.3 - Schematic sketch of influencing factors in the observed concentration profiles

(from Kunkel et al., 2004). However, if the evaluated monitoring data contain both natural, i.e. unaffected, and influenced groundwater samples, the observed concentration profile consist of a superposition of these two components. In this case the observed concentration distribution (fobs) can be described as the sum of two distribution functions, representing the natural and the influenced component (see figure 6.3):

( ) ( ) ( )cfcfcf inflnatobs += (2) Evaluation of the observed concentration profile allows a separation of the two components and the identification of both the natural and the influenced component. However, the shape of the two distribution functions is not known from a priori. An analysis of the observed concentration patterns using statistical tests show that both component distributions can be satisfyingly expressed by lognormal distributions (Kunkel et al., 2002). In this case the explicit shape of both distribution functions is determined by three independent parameters each (amplitude, median and variance), which have to be fitted to the observed frequency distribution using standard algorithms. Characterisation of natural groundwater conditions as well as the effects of anthropogenic impacts can be done by any two independent parameters characterizing the distribution function of the components. For the lognormal distribution, median and variance are the most common parameters given. However, these parameters don’t give a very transparent measure to specify the groundwater conditions. Therefore, the distribution is characterized by a concentration range defined by the 10% and the 90% percentiles of the concentration distribution. • Bulgaria



In Bulgaria, a statistical approach aiming to assess NBL has been developed. It is based on the study of three types of distributions (normal, lognormal, or exponential). The statistical approach has been applied to each of the 11 types of aquifers (see following paragraph for more details). For this work, data influenced by human activities and samples taken near ore deposits and tectonic faults have been excluded. The statistical proceeds showed that for major elements, the distribution is normal (in homogeneous rocks) or lognormal. For trace elements, the distribution is often exponential because many results are lower than the detection limit. In case of normal distribution, NBL is considered as being the median (or mean) value. • France The French Ministry of the Environment (Water direction) and the French Water Agencies entrusted the BRGM with the development of a methodology, applicable to the French geological context, able to characterise the natural baseline of water chemistry. A large range of factors controlling the chemical composition of groundwater overlap, making impossible a deterministic approach that could be applied to all types of aquifers. The proposed methodology therefore adopts a step-by-step strategy that takes into account the definition of groundwater bodies, the need to identify human pressure on the groundwater resources as well as the long-term quality objectives, fixed by the Water Framework Directive for groundwater bodies. To make sure that the methodology is practicable, it has been tested on four different case studies. All these results (state of the art, methodology and application to case studies) are compiled in a guidance document (BRGM, 2006). This document aims to give to water managers a practical methodology in order to assess the geochemical background of one given aquifer. It aims to be used by the River Basin Districts to meet the Water Framework Directive requirements. As the future Groundwater Daughter Directive will ask Member States to define threshold values for the definition of good chemical status, this guidance document offers key elements to distinguish anthropogenic inputs from naturally occurring substances. Before developing a methodology to assess the geochemical baseline of an aquifer, a state of the art about the knowledge of the natural chemical composition of the groundwaters has been realised, concerning: - major elements; - trace elements; - natural processes which may affect the chemical composition of a groundwater body. The state of the art on natural major elements occurring in groundwaters has been performed following two steps. First of all, a review of the literature dealing with natural background levels in French aquifers has been realized. This work gives for each type of aquifer the main attended elements. Next, a statistical analysis has been performed to define, for each type of aquifer a typical chemical composition. Only main classical statistics (min, max, mean, median, percentiles) have been calculated in order to obtain the typical profile of each type of aquifer. The typology of aquifers has been realized following 3 levels of criteria: - the lithology of the aquifer: chalk, crystalline basement, limestone, shales, sands and

sandstones, basalts);



- distance to the coast (< or >100km). The concentration of some major elements such as Cl and Na is indeed widely influenced by the distance to the coast. For these elements, the concentrations in rainfalls are indeed higher by the sea than inland;

- the oxydoreduction conditions: confined/unconfined. As an example, the following table shows the statistics for an unconfined limestone aquifer situated far from the seaside (table 6.1).

Table 6.1 - Example of statistics provided in France for each type of aquifer. Example of an unconfined limestone aquifer far from the coast.

As for major elements, the state of the art on trace elements includes an exhaustive review of the relevant literature as well as a statistical analysis (see above description). However, the statistical results are for trace elements less relevant than for major elements. Due to heterogeneous and sometimes high limits of quantification, the results are not always of good quality. Furthermore, the criterion used to avoid samples influenced by human activities (NO3 concentrations) is not always sufficient to remove the contaminated samples. From the review’s results (that covered French studies as well as foreign literature), the most favorable geological contexts have been defined for each parameter. • Lithuania In Lithuania, the statistically upper value (95% or 99%) is generally considered as the NBL. Although no national and official methodology exists, a statistical approach is proposed and has been recommended by the Minister of the Environment to be used for the initial characterisation process required by the WFD. This approach proposes to define the hydrochemical background as the average amount of element established by observing its natural variability in a natural object, which is homogeneous from the hydrogeochemical point of view. Its value depends on the selected estimation of the central (arithmetical mean, geometrical mean, logarithmical mean, median) value, of the probability distribution model and homogeneity of the samples. Under normal low, background distribution shall be distribution of the concentration of elements within the

interval ⎥⎦⎤

⎢⎣⎡ +−

−−

YY SYSY 96,1;96,1 , where −

Y is the estimation of the central position of data

(average), YS is the estimation of standard deviation Yσ , and values outside this interval



shall be regarded as abnormal. When the available data or their (transformed) logtransformed values correspond to normal distribution, 95% of the values of the samples fall in this interval. The statistically validated upper background limit of the macro-component composition of groundwater corresponds to the statistically validated (p≥5%) interval of

distribution model ⎥⎦⎤

⎢⎣⎡ +

−

YSY 96,1 .

Quite common is to calculate upper boundary of background as ⎥⎦⎤

⎢⎣⎡ +

−

YSY 2 .

As mentioned before, in Lithuania, the approach to assess NBL is made of common principles that could be summarised by the following steps:

Figure 6.4 - Steps and common principles aiming to identify NBL of groundwater in Lithuania.

6.2.2.3. Database In Germany as in France, the statistical analysis has been performed thanks to a national database. In both cases, a selection of non influenced sites has been necessary. The main criteria for the selection are nitrate concentration (<10mg/L) and the ionic balance (<20% in France). Finally, 26,000 monitoring points have been selected in Germany with one representative sample for each and nearly 18,000 in France (see following maps).

Verification and validation of data set

Calculation of statistical parameters

(min, max, STD, average, quartiles, percentiles)

Hydrochemical analysis (special diagrams, equilibrium equations)

Graphical analysis (Histograms of distribution of elements; peak analysis)

Factor analysis

Additional steps



Figure 6.5 - Selected monitoring sites for the statistical analysis aiming to identify NBL

of groundwater in France and Germany. In Bulgaria, two different studies based on two different databases are available: The first one includes 12,000 samples taken from 1965 to 1985. Major elements such as trace elements have been analyzed. The results obtained with this database are meanwhile influenced by mining and farmer’s activities (especially for heavy metals). In a second study, 11,800 chemical analysis have been used. No monitoring point is now located in an ore field. Each sample has been linked to one of the 11 types of lithology and geology: - deluvial sediments, clayey sands and clays with different geological age; - marls, argilites, siltstones, etc.; - quartzites, arkose, sandstones, conglomerates, shales, schists, etc.; - flysch formations; - carbonate rocks – limestones, dolomites as well as the marbles in Rhodopi Mountains; - alluvial deposits and alluvial; - loess; - volcanic-sedimentary formations; - acid volcanic and metamorphic rocks; - middle acid volcanic and metamorphic rocks; - ultrabasic and basic rocks – volcanic and metamorphic. In Lithuania, upper background limit usually is determined based on one from three data sources: - long-term National groundwater monitoring data; - data from temporary sampling of productive wells; - special case studies in which determination of NBL is one of the topics. The National groundwater monitoring network of Lithuania is made of 188 wells. Main elements are analysed once a year (usually in spring) and trace elements are measured twice every five years. The national network exists since 1963 and has been updated in 1995 and 2001. Lithuania underlines that data from this network can be considered as validated



from 1995 when quality assurance was introduced to monitoring procedures. Data collected for major compounds (Na, K, Cl, SO4, HCO3, CO3, Ca, Mg, NH4, NO3 and Fe) can particularly be considered of good quality. About trace elements (Pb, Zn, Cu, Mn, Mo, Ba, Al, F, Br) and organic compounds, less information are available. For these elements, laboratories have just recently been able to detect concentrations (limits of quantification) lower than drinking water standard. In this network monitoring shallow groundwaters, all the type of aquifers are monitored (sands, clayey deposits, gypsiferous deposits, rich in organic compounds materials). The following table presents an application of the statistic approach summarised above.

Table 6.2 - Chemical composition of shallow aquifers in Lithuania (mean value in 2003; 25% and 75% quartiles for the 1995-2003 period).

6.2.3. Local scale approaches (UK, DK, BL-Flanders, EE, NL, HU, PT) Many countries (UK, DK, BL-Flanders, EE, NL, HU, PT) did not develop a national approach to identify NBL and refer to case by case studies. Most of them contributed to the Baseline



project (see the following section for more detailed information) and propose to use this approach for the definition of NBL (UK, DK, PT, BL-Flanders, EE). During this project, NBL have been derived for 23 aquifers through Europe (cf. figure 6.6.).

Figure 6.6 - The 23 European aquifers studied during the Baseline project.

Denmark thus recommends considering NBL as a range of value and not as a single value. This concept was one of the main results of the Baseline’s project. Denmark summarizes the local approach to identify NBL in an aquifer as the following steps: - Application of environmental tracers and groundwater dating: If environmental tracers for

dating of young groundwater (e.g. 3H/3He, CFCs, 85Kr, SF6) are below detection limit, the groundwater composition is assumed to be close to the natural background level. An appropriate number of analyses from wells in such groundwater bodies will illustrate the natural variation of the analysed elements and compounds.

- Groundwater shows a human impact if young environmental tracers can be measured in the groundwater samples, which is the case for most Danish groundwater used for water supply. In this case estimation of the NBL is not easy, and a wide range of alternative methods are needed for evaluating the NBL for each relevant substance. These include: . evaluation of groundwater chemistry in similar hydrogeological settings where no

environmental tracers or contaminants have been measured, . evaluation of changes in hydrochemical composition with depth e.g. by sorting

monitoring data by depth. Such data and analyses show that human impact typically are found down to depth of about 80 m below surface in Denmark,

. evaporating the average composition of precipitation and equilibrating this with carbonate and other relevant substances during infiltration of water by e.g. reactive geochemical modelling,

. evaluating the general temporal and spatial hydrochemical evolution,

. evaluation of old data from the investigated or similar groundwater bodies,

. evaluation of time series of relevant substances from surface water (streams, lakes and coastal water),

. evaluation of lake sediment geochemistry and ecology.



Although some countries (Estonia, NL, Poland and Belgium-Flanders) don’t have any national approach and often prefer to realize a local scale study (aquifer by aquifer), they defined background values. In Netherlands, for example, the National Institute for Public Health and Environment (RIVM) deduced “semi-natural background levels” for 17 trace elements and in the three main types of geographical units of the country: sandy, peaty and clayey areas. The depth has also been taken into account. Three levels have thus been distinguished: 1m, 10m and 25m below surface. The semi-natural background levels were determinated as the 90 percentile plus its uncertainty. As these levels are calculated with the monitoring networks results, they can be influenced by human inputs (agricultural activities in sandy areas). In 2003, this work has been completed with the study of 50 other elements (major and trace) in sandy groundwaters. Data comes from national networks as well as from a local study (Veluwe area) where the determination of groundwater’s age allows considering this system as representative of natural groundwater. From these results NBL have been calculated (concentration in historic rain * evaporation factor). TNO (2004) deduced so-called geographically specific target values for 7 groundwater units in the western Netherlands (provinces of Northern and Southern Holland) for 7-8 trace elements. The groundwater units were distinguished on salinity and geographical location; here, three have a similar range of groundwater compositions. All kind of historic analyses were considered from screens having various depths, and anthropogenic influence is assumed to be minimal based on sampling depth and reactivity of shallow subsurface containing majorly clay and peat. The criterium for the target value was the 50 percentile or the 90 percentile minus its uncertainty.

6.2.4. The Baseline project This section aims at giving a summary of the content and of the main results of the Baseline project. This project, coordinated by the British Geological Survey (Mike Edmunds), involved the participation of 14 partners from UK, Denmark, Spain, Portugal, Belgium, France, Estonia, Poland, Switzerland, Czech Republic, Hungary and Malta. The work within this project was organized in the following workpackages: Workpackage 1: Overview of baseline concept A definition of baseline was formulated as follows: “the range of concentrations of a given element, species or chemical substance present in solution, being derived from natural geological, biological, or atmospheric sources” Workpackage 2: Baseline inorganic quality in European aquifers Data bases of 21 reference aquifers across a wide range of European conditions were compiled, as model examples of a recommended methodology for aquifer characterisation, and to illustrate the ranges in natural groundwater quality as a basis for understanding natural groundwater quality evolution and the interfaces with modern, probably contaminated water. Workpackage 3: Baseline organic quality in European aquifers The importance of the parameter TOC was stressed: at one hand because of its influence on the groundwater redox status, and on the other hand, as it may indicate pollution; however TOC-baseline may be high as well.



Workpackage 4: Geochemical modelling: geochemical controls and prediction The hydrochemistry of selected reference aquifers was modelled with PHREEQC. Workpackage 5: Timescales and chemical tracers as indicators of water quality change Isotopes and chemical tracers are important tools for understanding natural changes, and to detect the contribution of a young, possibly anthropogenically influenced component. Workpackage 6: Baseline trends in European aquifers Trends do not always reflect pollution. Besides pollution, trends may be due to fully natural changes at one hand, and may result from groundwater exploitation on the other hand (quantity changes leading to quality change). Also in the latter case, there is no question of an anthropogenic direct input of pollutants. Next to time trends, it is important to consider spatial trends; both are often related. Workpackage 7: Policy and end-users A questionnaire was sent to groundwater managers and end-users (water supply companies) in order to collect information on perception of groundwater-issues, and monitoring practices. Workpackage 8: Public awareness / data dissemination A flyer, aiming to inform and sensibilize the broad public on the value of the natural groundwater quality, has been prepared. Workpackage 9: Recommendations for monitoring practice Baseline monitoring should be focused at collecting data that contribute to system understanding, and to comprehend the natural variations of quality within the groundwater body, both in time and in space. This should serve as a reference, against which superimposed pollution can be detected. The main conclusions (after Edmunds, 2003: BASELINE project report) are: - The baseline is likely to be a range of values, varying both in time and space; - There is a need for a geochemical approach; the baseline is based on water-rock

interactions and natural cycles; - Reactions are time dependent and residence times are important; - Past and present environmental conditions need to be known; - Statistical treatment may be necessary but only as an aid to geochemical basis. About baseline and pollution, authors highlight that: - Pollution by definition is anthropogenic and superimposed on a baseline in space and

time; - Pollution signature may be recognisable by several unique chemical or isotopic tracers

and indicators. A number of substances intuitively considered pollutants (e.g. Cl, NO3, NH4, K, F, As and most metals) may have high natural baseline values under certain geochemical conditions.

Important recommendations are finally adressed: - A geochemical basis coupled with flow dynamics is needed for understanding groundwater

bodies; geochemical modelling provides a qualitative and quantitative understanding of the time-dependent, common major processes controlling groundwater quality (e.g. dissolution/precipitation, ion exchange, redox processes);



- Multiple tracers and dating tools allow baseline concentrations to be related to groundwater residence times (“ages”);

- A few of the reference aquifers (using state of the art tracers) demonstrate timescales and mixing in detail, serving as a basis for wider application of low cost indicators (e.g. chemistry, CFCs);

- Old waters (e.g. >100 years) are likely to be contaminant-free waters; - Baseline trends are generally related to disturbance of natural “fronts” due to pumping; - Monitoring should recognise and acknowledge the fact that flow conditions and

groundwater chemistry vary significantly laterally and vertically; - Groundwater quality is important in maintaining good surface water status. Surface water

is most times an “outcrop” of groundwater; hence surface water quality will be affected by groundwater quality;

- There is a need to promote the high purity of (most) natural groundwaters; - There is a need to consider quality as well as yield focus in licensing.

6.2.5. Conclusion About ten countries have return information about how they consider and assess NBLs in their countries. It appears that some of the countries developed a national approach in order to support groundwater managers to identify NBLs and to make the distinction between natural occurrences and anthropogenic occurrences (e.g. France and Germany). In other countries, no national approach has been performed but some local case studies have allowed characterizing NBLs in some aquifers (DK, UK, BL-Flanders, EE, PL, PT…). In the light of these first results, it seems important to keep in mind that: - in many countries, many efforts have been performed to identify NBLs even before the

publication of the WFD; - the level of knowledge meanwhile remains very variable from a country to another. When

Member States will apply the methodology, the “starting point” will be different. - the approaches are very different from a country to another. It is hence necessary to find a

practical approach that takes into account of all the available knowledge. 6.3. Typology of aquifers

6.3.1. Description of the typology In order to describe the natural background levels of each type of aquifers, a typology is proposed. It is based on two levels of criteria which are summarized in the table 6.3. This typology is presented here in order to remind that the NBLs are the result of very complex and numerous factors. These criteria won’t be detailed here (in particular the “further criteria”). The present chapter will only focus on the assessment of the “typical” chemical composition of each type of aquifer and will mainly refer to the prime criteria. From this typology, a description of the typical chemical composition (major elements and trace elements) of each type of aquifer is presented in the next two sections.



Table 6.3 - Typology of aquifers.

Limestones (karstic/not karstic) Chalk

Sands and gravels Sandstones Clays, marls

Crystalline basement rocks Volcanic rocks

Lithology

Schists, shales

Prim

e cr

iteria

Saline influence Hydrodynamics Recharge, residence time, topography, leakage...

Redox conditions Particular

occurrences Organic matter, oxide minerals… Furth

er

crite

ria

Geological age

6.3.2. Approach to define the “typical natural chemical composition” of each type of aquifer

The following sections aim to give a general overview of high natural background levels through Europe and to give some responses to the following questions: which elements will be present with high NBL? Where? Which range of values? In which geological context? It aims to give the “typical chemical composition” of each type of aquifer (for main and trace elements). It is meanwhile obvious that each aquifer is different and that the chemical composition of an aquifer is the result of very complex and numerous processes (interactions with the rock, flows, redox conditions, special occurrence like organic matter...). Aquifers are very complex and heterogeneous systems and for a precise identification of NBLs in each one, detailed studies are necessary. Unfortunately, the working scale of the WFD (25 Members State, hundred GWB) and the timetable don’t allow leading that kind of study for every GWB. Some general characteristics can meanwhile be identified and could help the Member States to identify the excepted (or possible) range of values of NBLs in each type of GWB. Following these first level of study, more detailed studies could be provided in order to understand the origin of an element. Notice: by dealing with processes controlling the behaviour of each substance, the work package 1 of the BRIDGE project included in its report a chapter about the major composition of groundwaters (chapter 4 of the D7). In order to provide coherent deliverables within the BRIDGE project, the present chapter includes these results. • Major elements Obviously each aquifer has its own chemical composition which is spatially and temporally variable. But, as the lithology and the water – rocks interactions greatly influence the chemical composition of the aquifer, it appears possible to provide the range of values of NBLs for each type of aquifer by using a statistical approach. Furthermore, it is important to note that this statistical approach also reflects some of the processes such as the residence time or the influence of the recharge. All these factors that lead to a particular chemical composition are included in the results used for the statistical analysis. As already mentioned in the section 1, some countries have performed a statistical analysis in order to give the range of values for each type of aquifers (FR, GE, BG, LT). These results are presented here in order to give the typical chemical composition of each type of aquifers.



Main statistics parameters are giver (min, max, percentiles, mean, median, standard deviations). Each one is important and provides different information (range of values, extreme levels…). Concerning the mean value, it’s important to note that it doesn’t really reflect the general composition of the aquifer. The median value is much more representative than the mean value. Notice: the following results mainly come from the French and the German approaches. The mains reasons for this are that: - the geological context of both countries is varied enough to provide information on all

types of lithology (see map figure 4.2); - the number of monitoring sites available allows to provide representative results (26,000

sites in Germany, 18,000 in France). Nevertheless, when it’s relevant and in order to complete the final issues, other results coming from Lithuania and Bulgaria will be discussed. For more detailed information about the methodology used by France and Germany to provide the statistical results, please refer to section 6.2 and to the appendix. It is meanwhile important to note that in spite of all the precautions that have been taken to provide representative data, the database used by France and Germany are not perfect and present a few limits: - first of all, even if a selection of the data has been realized (for example by only keeping

samples for which NO3<10mg/L), it is impossible to make sure that all the results are not influenced by human activities;

- for trace elements, due to heterogeneous and often high NBL, the statistics are not always representative. Furthermore, the selection criterion (NO3) is not sufficient to avoid all the point source pollution that could meanwhile influence trace elements concentrations.

Finally, although the results of the statistical approach can be considered as representative (especially for major ions), the results presented below are completed with information from the literature. This kind of information taken from reports or scientific papers is particularly important for trace elements for which, as explained before, the statistical analysis presents many limits (anthropogenic influence, high limits if quantification). Case studies presented in the literature allow to be sure of the natural origin of the measured concentrations. • Trace elements The following information on trace elements is based on the German statistical results and a review of the literature (that has been made by BRGM and completed by the contributions received during the project). This last approach is of major importance because: - trace elements analysis are not so current; - it is difficult to have data of good quality (low uncertainties, low limits of quantification); - the results can be influenced by the human activities (deterioration of the well, point

source pollution…) and the identification of the origin of an occurrence (anthropogenic or natural) require a detailed study.

Both approaches thus allow having a complete overview of natural occurrence of trace elements in groundwater.



6.4. Natural Background Levels in Limestones aquifers

6.4.1. Major elements • General comments When the dissolved constituents are derived primarily from dissolution of the rocks through which the water moves, the main constituents of carbonate rock aquifers (calcium, magnesium, and bicarbonate) are found in high concentrations (see the following tables). Meanwhile, without going too much in detail with the processes (which are not the main subject of this chapter), it’s important to remember that when identifying the NBLs other factors have to be taken into account. For limestones, the main criterion is the residence time. The greater the residence time the increased likelihood that the concentrations will be high. The karstic character appears as an important characteristic to be considered as it is tightly linked to the residence time and, as a consequence, it is illustrated separately in the following paragraphs. • Residence time As shown in the following figure, major ions concentrations generally increase from the recharge area to the deeper part of the aquifer (Na, Cl, SO4). As illustrated by the calcium levels, it is meanwhile not always the case. In this Jurassic limestone aquifer, we can thus observe a decrease in the calcium values, and a corresponding diminution of the chemically aggressive nature of the water with the increased distance covered beneath the confining layer. This example illustrates that water – rock interactions are very complex processes and even if some general rules can be identified (e.g.: increase of the concentrations with the distance covered), it remains very important to identify all the factors implied. • Karstic/non karstic It is necessary to distinguish karstic systems from non karstic aquifers. Compared to other carbonate rocks, the karstic aquifers display relatively greater depth to the groundwater table but at the same time a lower residence time, as groundwater tends to move through large open channels. Additionally at the same time, there is little contact area or contact time with the rocks. Hence, groundwater in karstic limestones may achieve flow velocities of a metre or more per second, rather than the centimetres per day typical for other rock units. The direct consequence is that non karstic aquifers generally present higher concentrations than the karstic aquifers (see following tables).



0

20

40

60

80

100

120

140

160

180co

ncen

tratio

n en

mg/

L

NaCaClSO4

Figure 6.7 - Evolution of the concentrations of major ions in a Jurassic limestones aquifer

(South-West France, Aquitaine Basin) (Blum et al., 2002).

Parameter N

Na mg/l 798

K mg/l 388

Mg mg/l 798

Ca mg/l 797

Fe mg/l 796

Mn mg/l 797

HCO3 mg/l 795

SO4 mg/l 798

Cl mg/l 798

NH4 mg/l 798

NO2 mg/l 807

NO3 mg/l 798

PO4 mg/l 596

DOC mg/l 190

LF µS/cm 871

O2 mg/l 861

H µg/l 873

pH -

10. P 50. P 90. P 1,3 2,9 6,3

0,33 0,80 1,9

3,0 22 37

67 97 125

0,004 0,03 0,15

0,000 0,001 0,003

296 329 366

13,0 20 32

4,6 13,1 37

0,0001 0,0001 0,01

0,002 0,004 0,009

0,83 3,8 17,5

0,01 0,04 0,14

399 567 706

4,8 9,0 10,8

0,02 0,05 0,08

7,1 7,3 7,7 Figure 6.8 - Natural background concentrations (major groundwater composition) in Carstic carbonate-rock aquifers in Germany based on ca. 1000 groundwater samples per parameter (Kunkel et al., 2004)

Flow direction Recharge Deep part (few hundred metres)

Unconfined Confined C

once

ntra

tion

in m

g/L



PParameter

N Na mg/l 748

K mg/l 640

Mg mg/l 753

Ca mg/l 753

Fe mg/l 739

Mn mg/l 704

HCO3 mg/l 749

SO4 mg/l 748

Cl mg/l 749

NH4 mg/l 705

NO2 mg/l 815

NO3 mg/l 753

PO4 mg/l 507

DOC mg/l 485

LF µS/cm 743

O2 mg/l 719

H µg/l 746

pH -

10. P 50. P 90. P 3,0 5,3 9,2

0,6 1,2 2,1

16,8 29 50

86 122 160

0,000 0,003 0,05

0,001 0,003 0,009

295 367 440

32 70 186

8,7 21 49

0,001 0,004 0,01

0,002 0,005 0,02

2,9 10,8 40

0,005 0,03 0,14

0,38 0,68 1,2

546 796 1105

3,0 7,6 10,2

0,02 0,05 0,08

7,1 7,3 7,6

Figure 6.9 - Natural background concentrations (major groundwater composition) in Carbonate-rock sequences in Germany based on ca. 750 groundwater samples per parameter (Kunkel et al., 2004).

• Magnesium With regard to magnesium concentrations, it’s important to remember that the type of limestone influence the level. Higher values are expected in dolomites. It is thus important to have a look at the main composition of the rocks when deriving the NBLs. This factor has been described and illustrated by the WP1 in its report. The following figure shows the observed distribution of the Mg – concentration in the German Jurassic limestones. It is shown here as an example for facies sequence differentiations, which have to be taken into account, when the major groundwater composition of the hydrogeological units is described.

Figure 6.10 - Distribution of Mg-concentration in the Jurassic limestone aquifers of Germany

(Kunkel et al., 2004). As shown in figure 6.10, the Mg-concentrations in the Jurassic limestone show a bimodal distribution. In the case of this hydrogeological unit the bimodality is neither caused by a possible high influence of anthropogenic intakes, nor does it represent an inhomogeneous



data set, i.e., groundwater samples from petrographical and/or hydrodynamical different hydrogeological units. Instead, the two components in the observed concentration profile are caused by natural reasons: a rising degree of dolomitisation in the Jurassic limestones from southwest to northeast.

6.4.2. Trace elements Although limestone aquifers are not particularly favourable to the presence of trace elements (see table below), it is important to note that specific occurrence can occur. The residence time and the redox conditions greatly influence these occurrences.

Parameter N

Komponentensep.

10. P 50. P 90. P Fe mg/l 796

Mn mg/l 797

0,004 0,03 0,15

0,000 0,001 0,003 F mg/l 402

Br mg/l 229

J mg/l 30

Ag µg/l 211

Al µg/l 334

As µg/l 374

B µg/l 192

Ba µg/l 401

Bi µg/l 260

Cd µg/l 470

Co µg/l 387

Cr µg/l 470

Cu µg/l 327

Hg µg/l 406

Li µg/l 385

Ni µg/l 471

Pb µg/l 470

Sb µg/l 260

Se µg/l 417

Si mg/l 762

Sn µg/l 260

Sr µg/l 394

Zn µg/l 428

0,04 0,07 0,14

0,95 2,1 4,8

0,19 0,49 1,3

7,6 15,8 33

0,007 0,04 0,23

0,10 0,18 0,32

0,15 0,49 1,7

0,36 0,60 0,99

0,005 0,02 0,06

0,28 1,2 5,1

0,24 0,48 0,95

0,01 0,10 0,70

0,28 0,58 1,2

27,6 60 131

0,70 3,1 13,8

Figure 6.11 - Natural background concentrations in karstic carbonate rocks in Germany (Kunkel et al., 2004)

6.4.2.1. Iron and Manganese

These two elements are the most current frequently found elements in limestone aquifers. Their occurrence is closely linked with the redox conditions and special occurrences (bacteria, organic matter...). As shown by the following figures, a reduced environment favours iron and manganese release.



For these reasons, in oxidized waters (a few mg/L of O2), the Fe3+ values don’t generally exceed a few tens µg/L. The concentration is limited by Fe(OH)3 precipitation. In reduced waters, the reduced form of iron Fe2+ is dominant and NBLs can be very high (a few mg/L). A low pH also favours iron occurrences. Examples of high NBLs for Iron and Manganese: - France: Jurassic limestones aquifer (SW of France, Aquitaine Basin): in this regional scale

aquifer, significant occurrences of iron (from 40µg/L to 3,060µg/L) and an increase of the Fe values has been observed from the recharge area to the deeper part of the aquifer situated 200km farther on.

As a conclusion, it is noted that the iron and manganese occurrences are not specifically linked to a type of lithology. Any reduced environment can present high NBLs of iron and manganese. This is also the case in a confined aquifer (according to a lack of sulphurs) and any other system where the oxygen has been consumed (in particular by organic matter degradation).

Figure 6.12 - Iron stability fields of the dissolved and solids phase versus pH

and Eh for 1 atmosphere and 25°C (Hem, 1985).



Figure 6.13 - Manganese stability fields of the dissolved and solids phase versus pH

and Eh for 1 atmosphere and 25°C (Hem, 1985).

0.0

0.1

0.2

0.3

0.4

0.5

0.6

0.7

0 20 40 60 80 100 120 140 160 180

distance d'écoulement (km)

Fe (m

g/L)

libre

captif

D. minéralisé

CaptifLibre

Limite 'chimique'

TOU SAV

Figure 6.14 - Iron concentrations evolution versus the distance covered by the water.

Jurassic limestone aquifer (SW France, Aquitaine Basin) (Blum et al., 2002).



6.4.2.2. Fluorine

In limestone aquifers, the main sources of fluorine for groundwaters are: - fluorspar CaF2: its solubility is relatively high; - fluorapatite Ca5F(PO4)3. A few mg/L is not a rare concentration for fluorine in groundwaters. The factors controlling the fluorine levels are (Travi, 1993): - the temperature and the pH which influence the solubility product; - the presence of complexes or colloids (MgF+ in particular); - the solubility of the minerals; - anionic exchanges (F-/OH- substitution). The F- ion is indeed easily adsorbed on some

clays (illite, gibbsite, kaolinite, aluminium hydroxides). Hence, fluorine included in the clayey surrounding beds can easily be released;

- calcium concentration: an inverse correlation between Ca and F levels exists; - sodium: it seems like in many groundwaters, a high Na concentration increases the

fluorspar solubility; - the time for water – rock interactions. For these reasons, large sedimentary basins (Paris basin, Aquitaine basin, London basin) can present high NBLs of fluorine. It particularly occurs where calcareous-marl layers are present. The presence of marine clays also favours the occurrence of fluorine.

0

1

2

3

4

F (m

g/L)

Figure 6.15 - Evolution of fluorine concentrations in a Jurassic limestones aquifer (SW France)

according to the covered distance (Blum et al., 2002). Examples - Lithuania: the highest natural concentration measured in Lithuania is 2.11mg/L in Permian

carbonates.

Libre Captif Domaine

minéralisé

Recharge Deeper part Flow direction



- France: In the Parisian and Aquitaine basins, the main calcareous levels in which high concentrations of fluorine have been found are: . Eocene: Lutetian limestones and marls near Paris (from 0.45 to 4.5mg/L), . Jurassic limestones in the Meuse area (mean value of 2.18mg/L in the Bajocian aquifer)

and in Normandy (with a maximum of 7.25mg/L in the Bajocian layers). The following figure shows the evolution of fluorine concentrations in a Jurassic limestones aquifer (SW France). It illustrates the importance of the time of water – rocks contact.

6.4.2.3. Arsenic

Arsenic doesn’t “typically” occur in limestone aquifers. Crystalline basement rocks are much more favourable to the presence of arsenic. Meanwhile, high NBLs of arsenic in limestone systems are possible. Like in any type of aquifer, high concentrations of arsenic are expected in layers crossed by sulphide veins. A reduced environment is also in favour of arsenic release. Finally, specific occurrences also control the arsenic concentrations (organic matter, arsenopyrite). Example of high As natural concentrations: - France: very high concentrations have been measured in the Bathonian and Callovian

limestones in the East of France. The values often exceed 100µg/L with a maximum of 6,263µg/L. The presence of numerous sulphide veins explains these values.

6.4.2.4. Selenium

Selenium is typical of sandstones and sandy aquifers. Meanwhile, high concentrations of selenium have been observed in limestones aquifers. For example in France, in many Jurassic limestones systems of the Paris basin, high NBLs have been observed (a few tens µg/L with a maximum of 60µg/L). As demonstrated by several local studies, these occurrences are the result of an upward leakage from the underlying Ypresian aquifers. The Ypresian sands of the Paris basin are indeed rich in organic matter (plant remains) which is a favourable context for selenium occurrence. Recently, increasing concentrations have been observed in different monitoring sites. Detailed studies are presently conducted in order to clearly identify the origin of this phenomenon. But it seems like overexploitation enforced the leakage process. This example reminds the importance of quantitative aspects on groundwater quality (see chapter 9).

6.4.2.5. Barium Barium is not very abundant in carbonates rocks (see table 6.4). Meanwhile, if an aquifer contains any barytine BaSO4 occurrences, it is possible to meet significant natural barium concentrations in limestones aquifers. In sedimentary environment, the witherite BaCO3 is another source of Ba. The barytine solubility is relatively low and the Ba concentrations are largely controlled by the SO4 concentrations. Hence, significant Ba levels require a low SO4 concentration (<10mg/L). Example: - France: Eocene of the Aquitaine Basin (from 121 to 546µg/L). These high values are local

and are associated to low SO4 concentrations (indicating their natural origin).



6.4.2.6. Aluminium Although Al doesn’t generally occur in carbonates rocks, some samples showed significant values. Clay layers are indeed an important source of Al and that’s why if a limestone aquifer contains clay layers, it can present significant natural concentrations of Al. Examples: - Lithuania: in the Devonian carbonates, the highest value analysed for Al is 2,290µg/L

Table 6.4 - Mean abundances (world-wide scale) of major and trace elements in the main types of rocks (ppm ou mg/Kg). From Horn and Adams (1966) in Hem (1985).

Sedimentary rocks Element Igneous rocks Sandstones Clays Carbonates Si 285,000 359,000 260,000 34 Al 79,500 32,100 80,100 8,970 Fe 42,200 18,600 38,800 8,190 Ca 36,200 22,400 22,500 272,000 Na 28,100 3,870 4,850 393 K 25,700 13,200 24,900 2,390 Mg 17,600 8,100 16,400 45,300 P 1,100 539 733 281 Mn 937 392 575 842 F 715 220 560 112 Ba 595 193 250 30 S 410 945 1,850 4,550 Sr 368 28 290 617 C 320 13,800 15,300 113,500 Cl 305 15 170 305 Cr 198 120 423 7.1 Cu 97 15 45 11 Ni 94 2.6 29 13 Zn 80 16 130 16 Pb 16 14 80 16 B 7.5 90 194 16 As 1.8 1 9 1.8 Sb 0.51 0.014 0.81 0.20 Hg 0.33 0.057 0.27 0.046 Cd 0.19 0.02 0.18 0.048 Ag 0.15 0.12 0.27 0.19 Se 0.05 0.52 0.6 0.32

6.4.2.7. Boron

Boron is not very abundant in carbonates systems (table 6.4). But as it is relatively abundant in marine clay layers it is possible to find significant occurrences of boron in limestone aquifers where the limestone is inter-bedded with marine clay layers.



In carbonates aquifers, the boron concentration can also be influenced by evaporates deposits such as colemanite Ca2B6O11.5H2O or kernite Na2B4O7.4H2O. Furthermore, as boron is relatively abundant in oceans, rainfall also represent a significant source of boron for aquifers situated near to the coast (<200km). Natural background levels for boron are very variable and a few tens or even a few hundred µg/L are not rare. But the main problem when interpreting boron concentrations is its potential anthropogenic origin. Boron is indeed a very common element in detergents and is often discharged in waters in urban areas. In order to distinguish natural boron concentrations from anthropogenic ones, it is possible to use isotopic tools such as 11B. Examples: France: different carbonates systems of present significant NBLs of boron: - the confined Bajocian aquifer in Normandy (from a few tens µg/L to 1.5mg/L); - the Dogger, Rhetian and Keuper units of the Paris basin for which evaporates deposits

influence explain these occurrences. In the Lutetian levels, very high NBLs of B have also been measured (up to 350mg/L). Once again gypsiferous layers explain these concentrations;

- Infra-Toarcian: in this clayey aquifer, high natural concentrations have been analysed. The highest value is equal to 1,168µg/L;

- Carboniferous limestones in the northern part of France (Orchies and Flanders basins). In the Artois-Picardie district, the boron concentrations naturally exceed 300µg/L in 20 sampling sites. Exchanges with marine clays explain these values;

- In this same basin, the Jurassic limestones around Boulogne often present significant boron concentrations (from 100 to 300µg/L). The origin of boron in this coastal aquifer would be the rainfalls (Pain, 1996).

6.5. Natural background levels in chalk aquifers

6.5.1. Major elements Figure 6.16 shows an example of the statistics obtained for the French chalk aquifers. As the influence of the coast and of the redox conditions appeared important, different results have been provided following these criteria.

6.5.2. Trace elements Like other carbonates rocks, chalk aquifers are not the kind of systems in which high NBLs are the most expected. Meanwhile, some elements, such as Ni, F, Fe, and Mn, can present high natural concentrations in chalk aquifers.

6.5.2.1. Iron and Manganese As already said before, high natural values are expected for these elements in any kind of lithology provided that the environment is reduced.



Craie ; LibreNO3 < 10 mg/l ; distance > 100 km de la côte

Paramètre T pH Cond Cond Eh O2 CO SiO2 Ca Mg Na K Cl SO4 HCO3 NO3 NH4 NO2 TAC P tot PO4

Unité °C µS/cm25°C

µS/cm20°C

mV mg/l mg/l mg/l mg/l mg/l mg/l mg/l mg/l mg/l mg/l mg/l mg/l mg/l °f mg/l mg/l

Code Sandre 1301 1302 1003 1004 1330 1311 1841 1348 1374 1372 1375 1367 1337 1338 1327 1340 1335 1339 1347 1350 1433

Nb de valeurs utilisées 98 152 35 126 5 119 7 148 154 154 154 154 154 154 154 154 154 154 123 107 64Minimum 8 5,5 122 81 53,2 0,2 0,45 2,4 5,2 0,5 2 0,1 3,6 1 14,6 0,1 0,01 0,01 1,2 0,02 0,01

1er quartile 12 7,2 458 405 90,6 1,2 0,5 9,05 81,3 2,4 5,5 0,95 11 10 257,42 1 0,02 0,01 21,3 0,09 0,03Médiane 13,2 7,35 550 471,5 170 2,45 0,52 14,725 92,55 4,325 8,675 1,7 15 19,4 289,7 1,8 0,05 0,01 24,15 0,2 0,1

3ème quartile 14,6 7,5 726 575 216 5,3 0,6 22 109 9,8 12 3,1 19,55 36,5 337 4,05 0,1 0,03 28 0,2 0,2Maximum 22 8,06 895 804 241 9,85 1,75 33 151 51 44,1 13 39,8 111,9 470 9,9 1,07 0,13 35,1 0,8 0,44Moyenne 13,63 7,29 557,06 489,75 156,64 3,34 0,70 15,95 92,36 7,85 9,80 2,37 15,83 26,26 290,70 2,94 0,10 0,03 24,12 0,17 0,12Ecart-type 2,56 0,35 195,72 129,83 63,24 2,64 0,43 7,59 28,43 9,43 6,36 2,09 7,36 23,12 83,71 2,80 0,17 0,03 6,56 0,12 0,10

Paramètre Fe B As Sb Pb Zn Se Ni Cr F Cu Mn Ba CommentairesUnité µg/l µg/l µg/l µg/l µg/l µg/l µg/l µg/l µg/l mg/l µg/l µg/l µg/l

Code Sandre 1393 1362 1369 1376 1382 1383 1385 1386 1389 1391 1392 1394 1396

Nb de valeurs utilisées 152 16 15 13 133 145 13 13 13 152 145 152 12Minimum 3 10 1 1 1 5 1 1 1 0,05 1 1 4

1er quartile 50 25 1 1 5 10 1 1 1 0,09 5 5 5Médiane 290 28 5 1 5 20 4 5 5 0,11 30 11 19

3ème quartile 870 50 5 10 5 50 5 10 5 0,19 40 31 31Maximum 9200 72 5 10 20 450 10 10 5 1,55 100 656 50Moyenne 783,06 34,75 3,20 4,54 5,45 33,07 3,46 4,69 3,54 0,19 28,96 34,94 19,00Ecart-type 1365,88 15,74 1,94 4,33 3,55 44,67 2,65 3,87 1,87 0,22 27,20 83,71 14,45

Figure 6.16 - Example of statistics obtained with samples from French chalk aquifers

that are oxidized and far from the coast.

6.5.2.2. Nickel High natural background levels are very common in chalk aquifers. As the Ni ion is easily substitute with/for Fe or Mg ions in sulphurous minerals, it is mainly present in minerals such as: pyrite (Fe,Ni)S2 or pentlandite (Fe,Ni)9S8. It can also occur associated with hydroxides such as garnierite (Ni,Mg)6[(OH)6Si4O11]H2O or limonite (Ni,Fe)O(OH).nH2O. But more often, the pyrite oxidation is the main source of Ni for groundwaters (due to a decrease of the water table or due to the presence of nitrates for example). Examples: - France: the chalk aquifer in the Northern part of the Paris basin. Many samples of this

regional scale aquifer present high NBLs for Ni. The maximum value measured is 100µg/L. These values usually occur in the confined part of the aquifer where the Eh decreases (100 to 300mV) and allows the reduction of the oxides containing Ni. Furthermore, it has been demonstrated (Vallee, 1999) that denitrification processes occurring in this aquifer are coupled to pyrite and marcassite oxidation. These two minerals are very abundant in the Cretaceous chalk. These reactions are also coupled to an increase of SO4 concentrations. It is also important to note that in the most reducing part of the aquifer, the Eh leads to Ni precipitation (sulphurous minerals).

- United Kingdom: in the London basin, high NBLs for Ni have also been observed. The highest value measured is 300µg/L and the percentile 90 of the results is equal to 7µg/L (see annex).

- Denmark: in the chalk aquifer, significant natural levels for Ni have also been observed. The highest value is 42µg/L. The reducing conditions and the presence of iron sulphides also explain these values (Hinsby et al., 2003c). But as stated earlier, the water-table decrease frequently promotes the Ni release.



6.5.2.3. Fluorine

Chalk aquifers are not particularly favourable to the presence of F. But as already mentioned before, high values are expected in carbonates aquifers containing marine clays. Meanwhile, high natural concentrations of fluorine have been measured in this kind of aquifer. Examples: - France: Upper Cretaceous (Senonian and Turonian chalk): 0.3 to 2.5mg/L.

6.5.2.4. Mercury Mercury rarely occurs with significant concentrations in natural groundwaters. The values are usually situated around 0.1µg/L. The main mineral source of mercury is the cinnabar HgS. Although, mercury is essentially abundant in igneous rocks and clays (table 6.4), significant concentrations in chalk aquifers have been measured in Denmark and United Kingdom. NBLs for mercury yet don’t generally exceed drinking water standards (1µg/L). Examples: - Denmark: in the chalk aquifer around Copenhagen, the highest value measured for Hg is

0.33µg/L. - United Kingdom: the highest value observed for Hg in chalk aquifers is 2µg/L but the 90th

percentile of the data is equal to 0.05µg/L thus reinforcing the understanding that the Hg NBLs are very low.

6.5.2.5. Aluminium

Aluminium is relatively abundant in igneous rocks, clays and also in sandstones. In carbona rocks it is not very abundant (table 6.4). The Al ion can substitute to Fe or Si in feldspar rich minerals and amphiboles. The Al ion frequently forms hydroxides such as gibbsite Al(OH)3 which is the most current aluminium hydroxide. Highest NBLs for Al are hence expected in crystalline basement aquifers and sedimentary aquifers containing clay layers or in sandstones aquifers. A concentration of a few tens µg/L of Al is not rare in groundwater. Even higher values are possible in acid or basic environments. Although, considering Al properties, high NBLs are not expected in chalk aquifer, it is possible to find around a few tens µg/L of Al. Examples: - Denmark: significant values have been observed in the chalk aquifer (mean value =

81µg/L; max = 5,740µg/L). These values are meanwhile below the generally observed values in other geological contexts in the Danish groundwater monitoring networks.

6.6. Natural background levels in sands and gravels aquifers

6.6.1. Major elements The glacially deposited unconsolidated sediments north of the low mountains are combined in the hydrogeological reference unit "Sands and Gravels of the North European Plain”.



Glacial deposits consist mostly of clay, silt, sand and gravel in various combinations, but also include cobbles and boulders. The general term "glacial drift" is used for all types of glacial deposits, regardless of the particle size or the degree of sorting of the deposits, or how the deposits were emplaced. The glacial drift was deposited during several advances and retreats of continental ice sheets. Glacial ice and meltwater from the ice laid down several types of deposits. Till, which is an unstratified, unsorted mixture of material that ranges in particle size from clay to boulders, was deposited under the ice or directly in front of the ice sheet. Likewise tills are no productive aquifers. Outwash deposits, by contrast, generally consist of stratified sand and gravel that form productive aquifers. Most of the outwash deposits are in valleys; the intervening hills are mantled with till. Before or during the Pleistocene, some rivers in the glaciated area cut their channels as much as 300 feet deeper than their present riverbeds. In general, these channels appeared in front of the Pleistocene ice sheets which moved from northeast to southwest. Consequently streams like the river Elbe or the river Weser typically flow from southeast to northwest and display deep so called ancient river valleys. Some of the deeply cut meltwater stream valleys were later filled to their present levels with glacial deposits and alluvium. Today they represent lowland regions close to the water table. Shallow Holocene deposits of marine, deltaic and fluvial origin are present in the transition zone to the seas (Baltic sea, North sea) and the river mouths. In these areas seawater intrusion might typically occur. Sand and gravel deposited as alluvium along the valleys of major streams also form productive aquifers. Some of the alluvium consists of reworked glacial deposits that were eroded and transported downstream during and following the last retreat of the ice. Most of the productive aquifers in the surficial aquifer system consist of valley-fill deposits of coarse-grained glacial or alluvial deposits, or both, and contain water under mostly unconfined conditions. In the primary statistical analysis of the German groundwater data performed prior to the evaluation of natural background levels it was found, in actual fact, that no significant differences in substance concentrations occur if a classification according to chrono-stratigraphic criteria was made. Hence, groundwaters from different glacification periods (Saalian, Weichselian) displayed the same composition. In case of a differentiation according to groundwater sampling depth, however, differences in the distribution patterns of the substance concentrations occurred. For this reason, the "Sands and Gravels of the North European Plain" were evaluated according to the sampling depth. Here, 3 depth ranges were differentiated: - sampling depths of less than 10m; - sampling depths between 10 and 25m; - sampling depths between 25 and 50m. Comparing the results of the 3 withdrawal depths it becomes clear that the concentration patterns are quite similar. Compared to the other hydrogeologic units (see following chapters) the general solution content of around 1,000 (µS/cm) reflects the petrografic as well as the hydromechanical conditions, i.e. intergranular porosity allowing a good dissolving of minerals and a long groundwater residence time due to shallow hydraulic gradients. For many parameters (e.g. Na, K, SO4, Cl) the solution content decreases with increasing depth. This is striking on first sight as in general it can be expected that the solution content will rise with increasing depth due to the longer residence times of groundwater. This behaviour can be explained however by the ubiquitous input of anthropogenic substances from the surface (e.g. fertilizers) with the percolation water.



Parameter N

Na mg/l 2019

K mg/l 2002

Mg mg/l 3307

Ca mg/l 3337

Fe mg/l 2097

Mn mg/l 2809

HCO3 mg/l 3369

SO4 mg/l 3387

Cl mg/l 3533

NH4 mg/l 3154

NO2 mg/l 1050

NO3 mg/l 3034

PO4 mg/l 1705

DOC mg/l 2720

LF µS/cm 2237

O2 mg/l 951

H µg/l 3401

pH - 3401

Komponentensep.

10. P 50. P 90. P 6,9 16,2 38,1

1,0 1,9 3,8

3,4 8,7 22,2

27,1 71 153

0,06 0,7 8,0

0,06 0,3 1,4

28,9 150 351

7,0 36,4 189

12,2 32,6 87

0,011 0,03 0,1

0,004 0,02 0,06

0,1 0,3 0,9

0,01 0,04 0,11

1,0 3,0 8,8

226 474 993

0,2 1,0 6,0

0,01 0,03 0,2

6,8 7,5 8,2

Figure 6.17 - Natural background concentrations in Pleistocene gravels and sands in Germany (sampling depths <10m) (Kunkel et al., 2004).

With regard to the redox status, the distribution of the parameters O2, Fe(II), Mn(II), NO3, DOC show, that in general the hydrogeologic unit Sands and Gravels of the North European Plain can be regarded as carrying reduced groundwater. Due to the long groundwater residence times and high redox reactivity of the subsurface, a complete use of the oxygen reaching the aquifers with the percolation water, is attained. An indicator for this is the decreasing DOC and the fact, that nitrate is present in very few samples only, while there is a general trend towards high Fe(II) and Mn(II) concentrations. The absent nitrate is due to denitrification processes by which nitrate inputs to the aquifers is reduced to molecular nitrogen by microbially controlled redox reactions (associated with an increase of SO4 or alkalinity). If a groundwater is largely free of dissolved oxygen, certain micro-organisms are able to satisfy their oxygen demand by reducing nitrate. An important prerequisite for this reaction is the presence of organic carbon compounds and/or pyrite (FeS2) in the aquifer acting as reducing substances. This hydrochemical groundwater condition is typical for the unconsolidated glacial sand and gravel deposits of the North European Plain. Due to the denitrification processes in reduced aquifers described above, it may be possible that a groundwater appears to be almost free from anthropogenic nitrate inputs, although the nitrate input with the percolation water is in fact very high. Hence, the low nitrate contents in the loose rock sediments of the North German Plane are part of the specific hydrogeochemical conditions in the aquifers of this hydrogeologic unit.



6.6.2. Trace elements

6.6.2.1. Iron and Manganese Examples of high NBLs for Iron and Manganese: - Aptian sands aquifer of the Paris basin: in this confined part of the aquifer, high natural

concentrations of iron have been measured (up to 12.89mg/L). A low pH and a reduced environment emphasized by organic matter degradation explain these high values. In this same system, the Mn levels are also high: mean values around 87 and 203µg/L according to the season and a maximum of 1,328µg/L. Once again, the gentle acidity of the pH (6.55 to 6.88) and a decrease of the Eh due to oxygen consumption by organic matter oxidation explain these results.

- Denmark: in the quaternary sands, high natural iron concentrations have been measured (maximum of 17mg/L). In this aquifer, the manganese concentrations are also high and the highest value observed is 1.9mg/L.

- Finland: in quaternary sediments, important iron and manganese concentrations are not rare in Finland. Using the results of the national groundwater monitoring network, Soveri et al. (2001) thus report a mean value for iron of 189 µg/L for sands and gravels aquifers (2784 samples from around 50 sampling sites) and of 487 µg/L for moraine aquifers (2096 samples from around 50 sampling sites). Concerning manganese concentrations, the mean values observed on the same monitoring sites are 20.5µg/L for sands and gravels aquifers and 35.4µg/L for moraine aquifers.

6.6.2.2. Arsenic

As explained before, in sedimentary aquifers, arsenic occurrences mostly depend on sulphides veins and redox conditions. Therefore, in sandy aquifers, high natural values are not really expected for As but as it has been illustrated by several case studies, in reduced conditions, a concentration around a few µg/L or even a few tens µg/L is possible. Examples: - Belgium (Flanders): in the Neogene sands deposits, significant concentrations of As have

been measured. Out of 38 samples, the median value is equal to 6µg/L and the 95% percentile = 47.3µg/L. In some of the catchments, an increase in arsenic concentrations has been observed and is linked to groundwater overexploitation (Coestiers et al., 2005; See chapter 9 for more detailed information).

- Denmark: in the meltwater sands (Miocene sands + quaternary sands), significant values of As have been observed. The highest value measured is 46µg/L and the mean value = 4µg/L.

- Hungary: in the deep (>50m) reduced, and old (several ten thousand years ago) groundwaters of the Hungarian porous GWB, high natural arsenic concentrations have been measured. In the deeper layers almost 50% of the sampling sites present concentrations higher than the drinking water standard (10µg/L). Arsenic is present in groundwaters as As-oxihydrates. Redox conditions are drivers of the processes; adsorbed iron is needed to take arsenic in immobilised form, in contrary reduction of Fe- and Mn-oxihydrates (in reducing conditions and presence of organic matter) mobilize arsenic to be adsorbed on the surface of solid matrix. The most probable origin of arsenic is minerals of Transylvanian mountains. Groundwater flow regimes in the Quaternary have mobilised and transported arsenic. The accumulation process, multiplying the arsenic content up to 1,200µg/l, can be detected nowadays in shallow groundwater of discharge areas. Area of enrichment and accumulation has always been connected to the actual regional and local



discharge areas which have varied through geological ages, resulting the spread of arsenic into great part of sediments, by former groundwater flow regimes. Former discharge conditions are the reasons of high arsenic content detected in aquifers recently being recharge areas (due to change in topography in Danube-Tisza Ridge, Nyírség).

Figure 6.18 - Percentage of arsenic content higher than 10µg/L in Hungarian GWB.

6.6.2.3. Boron

As in carbonates rocks, boron is not very abundant in sands and sandstones. Meanwhile, as it has already been explained, the main source of boron is marine clays. Hence, if a sandy aquifer is inter-bedded with clayey layers, high natural concentrations of boron can be expected. Examples: - Belgium (Flanders): the Eocene marine Ledo-Paniselian sediments which are made of

sands and clays intercalations present high natural background levels of boron. The investigations performed in this aquifer during the Baseline project (see Edmunds et al., 2003; contribution from Belgium-Flanders in appendix) showed that the median concentrations was relatively important (513µg/L; 20 samples) and that the values could sometimes be very high (maximum = 2,336µg/L; 95th percentile = 1,655µg/L).

6.6.2.4. Nickel

Although high Ni concentrations are expected more often in chalk aquifer, significant values can occur in sandy aquifers as it has been illustrated in France or in Belgium. Reduced conditions often explain these levels.

0%

20%

40%

60%

80%

100%

freq

uenc

y

B K

S (<

20m)

P1 (20

-50m

)

P2 (5

0-10

0m)

P3 (10

0-20

0m)

P4 (2

00-400

m)

P5 (>

400m

)

groundwater type

Percentage of groundwaters having Arsenic content higher than 0,01 mg/l

>0,01 mg/l As <0,01 mg/l As

K the fractured and fissured sediments of the mountainous regions B bank-filtered resources P groundwater of porous sediments subdivided by depth from the terrain: o S the shallowest (<20m) i.e. the most vulnerable groundwater o P1 20 to 50m, o P2 50 to 100m, o P3 100 to 200m, o P4 200 to 400m, o P5 >400m (thermal water).



Examples: - France: Aptian sands: a study shown that among 44 measures, 7 exceed 10µg/L with a

maximum value of 41µg/L (Bourg et al., 1987 in BRGM, 2006). The Ni release is here associated to iron and manganese oxides solubilization.

- Belgium (Flanders): In the Ledo-Paniselian aquifer made of clays and sands, significant natural concentrations of Ni have been observed (maximum = 10.8µg/L, see Belgium-Flanders contribution in appendix).

- Finland: significant Ni concentrations have been measured in Finnish quaternary deposits (Soveri et al., 2001). The 90-percentile of the gathered data is 4.48µg/L in clayey and till aquifer lying on amphibolites bedrocks.

6.6.2.5. Fluorine

As mentioned before, the main kind of lithology leading to high NBLs for F are clays and crystalline basement rocks. But as clay layers can be included in sandy aquifers, they can be a source of fluorine in groundwater. Examples: - Belgium (Flanders): in the Eocene marine Ledo-Paniselian sediments made of a

succession of alternating clayey and sandy Tertiary deposits, high natural backgrounds levels have been observed for fluorine (see Flanders contribution in appendix and “Baseline” project results, Edmunds et al., 2003). In some samples, the drinking water standards are exceeded (1.66mg/L).

6.6.2.6. Selenium

The selenium’s chemical properties are close to those for sulphur. It is therefore often associated with iron or uranium. The main mineral containing selenium is ferroselite FeSe2. Hence the main source of selenium for groundwater is sulphide oxidation. Selenium minerals are associated to uranium deposits and are thus expected in sands, sandstones or conglomerates with plants remains. Examples: - France: several sandy aquifers of the French sedimentary basins present high NBLs for

Se: . Eocene (middle and upper) in the Aquitain basin: in tow monitoring sites, with levels of

10 and 17µg/L, the concentrations exceed the drinking water standard (10µg/L), . The Ypresian sands contain plants remains and represent a source of Se for

groundwaters. - Belgium (Flanders): in the Eocene marine Ledo-Paniselian sediments (clayey and sandy

Tertiary deposits), significant natural background levels have been measured. The mean value (20 samples) is around 1µg/L and the highest value observed is 5.3µg/L.

6.6.2.7. Ammonium

Although ammonium occurrence is essentially linked to redox conditions, some examples of European aquifers made of sands and gravels containing high NBL of ammonium can be quoted.



It is the case of the porous GWBs of Hungary. Ammonium exceeds the 0.5mg/l drinking water standard in many wells situated in porous aquifers. Groundwater resources of more than 50m deep aquifers contain ammonium of natural origin (due to in situ NH4 production in the aquifer or upward migration of NH4 of high depth where ammonium is resulted from early maturation of semi-maritime organic substances). This former origin is dominant in porous and karst thermal waters.

Figure 6.19 - Percentage of samples for each type of Hungarian GWB

in which NH4 concentrations are higher than 0.5mg/L. Extremely high ammonium concentrations in shallow aquifers (<20m) are caused by transport of natural NH4 upward in discharge area. The appearance of anthropogenic ammonium (originating from pollution of agriculture, settlements and animal farms) can not be excluded in recharge or in neutral areas (surplus of recharge and discharge is varying in function of meteorological conditions, no long term dominancy). Oxidising conditions in the unsaturated zone are generally enough for nitrification of all ammonium which was not taken up by plants. In this case NH4-content is negligible. If nitrification by bacteria is not total (small residence time in the unsaturated zone and reducing condition close to the groundwater table in discharge areas with shallow groundwater table) the rest of ammonium reaches the groundwater.

K the fractured and fissured sediments of the mountainous regions B bank-filtered resources P groundwater of porous sediments subdivided by depth from the terrain: o S the shallowest (<20m) i.e. the most vulnerable groundwater o P1 20 to 50 m, o P2 50 to 100 m, o P3 100 to 200 m, o P4 200 to 400 m, o P5 >400 m (thermal water).

0%

20%

40%

60%

80%

100%

freq

uenc

y

K (cold

) B

S (<20

m)

P1 (20

-50m)

P2 (50

-100m

)

P3 (10

0-200

m)

P4 (20

0-400

m)

P5 (>4

00m)

K (therm

al)

groundwater type

Percentage of groundwaters containing more than 0,5 mg/l ammonia

<0,5 mg/l NH4>0,5 mg/l NH4



6.6.2.8. Mercury The lack of references on mercury in groundwater doesn’t allow defining and characterising its natural presence in groundwater and its favourable geological context. The main mineral source of Hg is cinnabar (HgS). This mineral is generally present in crystalline basement rocks. Examples: - Finland: whatever the geological context, Hg concentrations are very low (median = 0.01

µg/L and 90-percentile = 0.03 µg/L). Natural Hg concentrations are thus very low. - France: the highest natural value known is a 0.06 µg/L value measured in the mio-

pliocene sands of the Landes (S-W of France). 6.7. Natural background levels of sandstones aquifers

6.7.1. Major elements

6.7.1.1. Triassic sandstones Aquifers in sandstones are more widespread in Germany than those in all other kinds of consolidated rocks. Although the porosity of well-sorted, unconsolidated sand may be as high as 50 percent, the porosity of most sandstones is considerably less. During the process of diagenesis of sand into sandstone (lithification), compaction by the weight of overlying material reduces not only the volume of pore space as the sand grains become rearranged and more tightly packed, but also the interconnection between pores (permeability). The deposition of cementing materials such as calcite or silica between the sand grains further decreases porosity and permeability. The average intergranular porosity of sandstone aquifers generally doesn’t exceed 10%. Thus, most of the porosity in these consolidated rocks consists of secondary openings such as joints, fractures, and bedding planes. Groundwater movement in sandstone aquifers primarily is along bedding planes, but the joints and fractures cut across bedding and provide pathways for the vertical movement of water between bedding planes. Because the number of joints, fractures, and bedding-plane openings in all types of rock typically decreases with depth, permeability of the sandstone aquifer should similarly decrease until groundwater movement ceases at some depth. Decreased circulation of groundwater results in an increase of TDS in groundwater In Germany, sandstone aquifers are parts of complexly embedded sequences of various types of clastic sedimentary rocks, predominantly from the Triassic epoch. The sandstones display a red colour, which is due to the arid-continental deposition conditions and gave this epoch its name (Bunter Sandstone comes from the German for "coloured sandstone", where it is now known as “Buntsandstein”). In general the sandstone aquifer system consist of layered rocks differentiated vertically into fine-grained, low-permeability rocks such as shale or siltstone, and more permeable predominantly sandstones. This unit is found in South and Central Germany, above all from Rhineland-Palatinate, Baden-Würtemberg, Hesse and Thuringia (see following figure 6.21).



Parameter N

Na mg/l 1617

K mg/l 1574

Mg mg/l 1629

Ca mg/l 1625

Fe mg/l 1552

Mn mg/l 1620

HCO3 mg/l 1604

SO4 mg/l 1622

Cl mg/l 1632

NH4 mg/l 1477

NO2 mg/l 1909

NO3 mg/l 1644

PO4 mg/l 439

DOC mg/l 453

LF µS/cm 1491

O2 mg/l 1599

H µg/l 1630

pH -

F mg/l 910

Br mg/l 196

J mg/l 116

Ag µg/l 378

Al µg/l 1376

As µg/l 669

B µg/l 522

Ba µg/l 542

Bi µg/l 195

Cd µg/l 1013

Co µg/l 335

Cr µg/l 1000

Cu µg/l 520

Hg µg/l 784

Li µg/l 363

Ni µg/l 1038

Pb µg/l 1019

Sb µg/l 305

Se µg/l 414

Si mg/l 670

Sn µg/l 210

Sr µg/l 390

Zn µg/l 684

Komponentensep.

10. P 50. P 90. P 1,9 5,5 16

1,3 2,1 3,6

1,9 6,5 22

5,0 11,3 26

0,002 0,02 0,09

0,001 0,004 0,07

6,4 25 95

5,3 18 58

4,0 8,3 17,4

0,001 0,004 0,01

0,002 0,004 0,009

2,2 7,5 26

78 237 692

4,9 8,1 10,8

0,03 0,07 0,19

6,7 7,1 7,6

0,03 0,07 0,2

2,2 9,3 39,0

0,3 1,0 3,1

6,2 11,8 22

26 95 350

0,007 0,05 0,41

0,14 0,53 2,0

0,32 0,91 2,6

0,03 0,05 0,08

0,05 0,48 4,4

0,002 0,04 0,75

0,29 0,50 0,87

3,9 8,1 17

Figure 6.20 - Natural background concentrations (major groundwater composition) in Sandstone aquifers in Germany based on ca. 1700 groundwater samples per parameter (Kunkel et al., 2004).

Composition



The sandstone sequences are in general characterized by a high significance for water management issues. Due to the siliceous nature of the rock, however, the waters frequency display low TDS and are threatened by acidification due to thin and buffer-poor soils. Thus more than 50% of all samples display an electric conductivity of less than 250µS/cm and a pH less than 7.0 (see table 6.3). Due to the relative high permeability of the sandstone aquifers and the steep hydraulic gradients, the residence time of the groundwater is generally shorter compared to the groundwater in the unconsolidated rock units. Groundwater recharge has been calculated to represent about 50% of the total runoff (Bogena et al., 2003), which is a good indication of the water-yielding capacity of the aquifer that provides the base flow. The Fe, Mn, O2 and SO4 concentrations show that the groundwater of the sandstone aquifers can be classified as predominantly oxidized. Thus, low Fe and Mn concentrations coincide with high O2 and NO3 concentrations. Although Ca is usually the major cation, Na and Mg are relevant as well. On milliequivalent basis, SO4 is the major anion instead of alkalinity.

6.7.1.2. Sandstone and clay (or siliceous clay) alternating sequences Observation wells that originate from alternating sequences of sandstone and clay from different ages were allocated to this hydrogeological unit. Especially observation wells from the “Rotliegend” as well as alternating sandstone and clay sequences from the Mesozoic, which do not belong to the Lower Triassic, were allocated to this unit. The latter contain as a rule, although subordinately, carbonate rocks and carbonate-cemented sedimentary rocks.

Parameter N

Na mg/l 1051

K mg/l 1040

Mg mg/l 1167

Ca mg/l 1170

Fe mg/l 1056

Mn mg/l 1006

HCO3 mg/l 1048

SO4 mg/l 1171

Cl mg/l 1192

NH4 mg/l 1117

NO2 mg/l 982

NO3 mg/l 1171

PO4 mg/l 720

DOC mg/l 344

LF µS/cm 1020

O2 mg/l 952

H µg/l 998

pH -

10. P 50. P 90. P 1,2 5,3 24

0,76 2,4 7,4

4,0 19,5 51

25 76 134

0,01 0,03 0,10

0,0001 0,004 0,07

60 280 403

14,2 37 95

5,2 16,8 55

0,0003 0,001 0,004

0,001 0,003 0,01

2,1 4,2 8,5

0,001 0,002 0,005

241 539 875

7,2 8,6 10,4

0,02 0,05 0,12

6,9 7,3 7,7 Figure 6.21 - Natural background concentrations (major groundwater composition) in sandstones and silicatic alternating sequences in Germany based on ca. 1,700 groundwater samples per parameter

(Kunkel et al., 2004).

6.7.2. Trace elements Several trace elements are expected to be present with high NBLs in sandstones aquifer: Se, F, Ba, As, Ni, Fe and Mn.



6.7.2.1. Iron and manganese

As explained in the above paragraphs, iron and manganese occurrences are not particularly linked to the lithology but mainly depend to the redox conditions. Hence, high NBLs can occur in sandstones aquifer if the environment is reduced. Examples: - France: in the Triassic sandstones of the eastern part of France, high natural background

levels of Fe have been measured (maximum value = 1.015mg/L). The reduced conditions in this confined part of the aquifer explain these concentrations (Eh = 10 to 185mV, pH≈7). The Mn concentrations are situated between 50 and 215µg/L).

- United-Kingdom: the confined part of the Triassic sandstones also presents high NBLs for iron and manganese. The highest values measured are 1,200µg/L for iron and 6,800µg/L for manganese (but with a 90th percentile for Mn equal to 270µg/L).

6.7.2.2. Fluorine

Examples: - France: in the Triassic sandstones of the Meuse region, the mean value of the F

concentration is equal to 1.87mg/L. - Estonia: the Cambrian-Vendian silts- and sandstones aquifer presents high NBL for F (see

Baseline project results, Edmunds et al., 2003 and Estonian contribution in appendix). The highest value is 1.14mg/L with a median value of 0.73mg/L (22 samples). As for barium, these values are due to the influence of the underlying crystalline basement aquifer and its weathering core.

6.7.2.3. Barium

In sandstones systems, barium is not very abundant but is a little bit more than in carbonates rocks (table 6.4). The main source of Ba in sandstones is also barytine BaSO4. Examples: - France: in the Triassic sandstones aquifer (Ardèche, SE France), high natural

concentrations have been measured (several samples >1mg/L; maximum = 1.6mg/L). These values are linked to the presence of barytine. This Ba rich environment also concerns the soil. Soils samples have shown high values of Ba. A positive correlation between the values in the soil and the concentrations in the groundwater has been demonstrated (see figure 6.22). In this aquifer, the correlation between Ba and SO4 also underlines the link between these two elements (figure 6.23). A significant Ba value requires a low SO4 concentration.



1

10

100

1000

10000

100 1000 10000

Baryum in soils (ppm)

Bar

yum

Con

cent

ratio

ns in

Wat

ers

(µg/

l)

Figure 6.22 - Relationship between soil geochemistry and concentrations in groundwater for Ba.

Triassic sandstones (Ardèche, France) (Barbier and Chery, 1995).

1

10

100

10 100 1000 10000

Baryum (µg/l)

Sulfa

tes(

mg/

l)

Figure 6.23 - Ba versus SO4 in Triassic sandstones aquifer (Ardèche, France)

(Barbier and Chery, 1995). Estonia: the Cambrian-Vendian silt and sandstone aquifer presents very high NBL for Ba (see Baseline project results, Edmunds et al., 2003 and Estonian contribution in appendix). The values can exceed 2mg/L. But it is very important to note that here, the origin of Ba is due to the underlying crystalline basement rocks and its weathering core. Groundwater in the clayey weathering core of the crystalline basement rocks is hydraulically connected with the underlying Cambrian-Vendian aquifer. Thus, overexploitation can enforce this phenomenon and lead to higher Ba values (see chapter 9).



6.7.2.4. Selenium As already discussed before, in sandstones aquifer with plants remains high natural background levels of selenium can occur. Examples: - France: in the Triassic sandstones of the eastern part of France, significant concentrations

of Se (maximum = 12µg/L; mean = 2.2µg/L) have been measured. But the origin of selenium in this aquifer is not yet clearly identified. A leakage from another Se rich system could be involved.

6.8. Natural background levels in clayey and marly aquifers As they can locally present relatively productive area, clays and marls can be considered as a type of aquifer. As mentioned several times in the other paragraph, clay layers can be a source for different trace elements such as Al, B, Fe, Mn, F, Zn. It is thus recommended to refer to other paragraphs to have information on these elements. Concerning major elements, as there is no specific issue to deal with, these won’t be studied here like in other sections. 6.9. Natural background levels in crystalline basement rocks aquifers Plutonic igneous and metamorphic rock sequences are summarized in the hydrogeologic unit crystalline rock aquifers. In general the spaces between the individual mineral crystals of crystalline rocks are microscopically small, few, and generally unconnected; therefore, porosity is insignificant. Many studies have determined that virtually all movement of water in crystalline rocks is through fractures or joints in the rocks. Fracture permeability in crystalline rocks is the result of the cooling of igneous rocks, deformation of igneous and metamorphic rocks, faulting, jointing and weathering. Openings commonly are present along relict bedding planes, cleavage planes, foliation, and other zones of weakness in the rocks. These openings typically are heterogeneous in spacing, orientation, size, and degree of interconnection. Generally, openings in the rocks are most prevalent near land surface and decrease in number and size with depth. Groundwater recharge to the crystalline-rock aquifers is at least less than 30% of the total runoff (Bogena et al., 2003). Movement of water through the rocks is totally dependent on the presence of the secondary openings; they generally yield only small amounts of water to wells, whereby the rock type has little or no effect on ground-water flow. Crystalline rocks generally are composed of virtually insoluble minerals, water is in contact with a relatively small surface area in the joints and fractures, and water movement generally is rapid and along short flow paths. Consequently, only small quantities of minerals become dissolved in the water.

6.9.1. Major elements The net effect of these factors is that the water contains little dissolved minerals and has an electric conductivity of less than 200µS/cm in most samples (table 6.#). As a consequence the TDS of groundwater in the crystalline rock aquifers is considerably lower compared to all



other hydrogeologic units. Due to the silicatic rock attribute and the thin buffer poor soils, groundwater is threatened by acidification. Thus more than 50% of all samples display a pH less than 6.5. The redox status for this unit is oxic, as illustrated by the high O2 concentration. Not only Ca, Mg and HCO3 are major ions as for the previously describe units, but Na and SO4 as well. The major presence of Na points out the importance of feldspar weathering.



Parameter N

Na mg/l 1025

K mg/l 1019

Mg mg/l 1045

Ca mg/l 1042

Fe mg/l 992

Mn mg/l 845

HCO3 mg/l 840

SO4 mg/l 1043

Cl mg/l 1046

NH4 mg/l 1032

NO2 mg/l 966

NO3 mg/l 1075

PO4 mg/l 721

DOC mg/l 189

LF µS/cm 894

O2 mg/l 1007

H µg/l 801

pH -

F mg/l 569

Br mg/l 491

J mg/l 69

Ag µg/l 439

Al µg/l 653

As µg/l 720

B µg/l 561

Ba µg/l 548

Bi µg/l 478

Cd µg/l 762

Co µg/l 509

Cr µg/l 731

Cu µg/l 108

Hg µg/l 728

Li µg/l 481

Ni µg/l 730

Pb µg/l 721

Sb µg/l 510

Se µg/l 553

Si mg/l 677

Sn µg/l 481

Sr µg/l 507

Zn µg/l 750

Komponentensep.

10. P 50. P 90. P 1,3 5,2 20,5

0,46 0,8 1,4

0,66 2,5 9,2

1,6 9,9 62

0,0004 0,007 0,11

0,002 0,004 0,07

4,9 16 55

1,7 12,0 83

0,03 0,60 14,2

0,001 0,003 0,01

0,001 0,003 0,01

1,5 5,1 16,9

0,003 0,012 0,05

0,32 0,88 2,4

27 74 206

6,4 9,0 11,0

0,03 0,29 3,0

5,5 6,5 7,6

0,74 3,6 18

0,0 0,3 2,5

1,1 2,4 5,0

0,0002 0,01 0,36

0,03 0,21 1,3

0,02 0,04 0,08

0,01 0,21 8,5

0,01 0,04 0,33

0,002 0,15 12,4

Figure 6.24 - Natural background concentrations (major groundwater composition) in crystalline rock aquifers in Germany based on ca. 1,000 groundwater samples per parameter (Kunkel et al., 2004).

Same comment as above.



6.9.2. Trace elements

In crystalline basement aquifers, it is no rare to meet high NBLs for many elements such as As, Sb, F, Al, Ba, Fe, Mn.

6.9.2.1. Aluminium As already said before, highest natural values are expected in crystalline basement aquifers (up to a few hundred µg/L).

6.9.2.2. Arsenic As already mentioned before, As can substitute to Fe and be present in sulphides such as arsenopyrite or in hydroxides. Arsenic can therefore be present in crystalline basement rocks and according to favourable pH-Eh conditions be released in groundwater. Examples: France: many cases with high natural concentrations of As are known in France: - Mont-Blanc and Aiguilles-Rouges: although the arsenic concentrations are spatially

heterogeneous, high natural background levels have been measured in these granites (almost 30µg/L; Dubois & Parriaux, 1990). In this area, the monitoring network aiming to check the quality of groundwater intended to provide drinking water also showed high arsenic levels. In other area of the French Alps, natural arsenic occurrences have also been identified (from 4.4 to 104µg/L in the Lauziere massif; a few tens µg/L in the Belledonne massif).

- In the Central part of France (Auvergne), several studied showed that in the granites and the gneiss areas high NBLs were expected for As (up to 140µg/L). These occurrences are linked to the abundance of arsenic sulphides in the hercynian granites.

- Bulgaria: 10-20µg/L of arsenic have been measured in crystalline basement aquifers.

6.9.2.3. Antimony Antimony is generally not common in groundwater but in some crystalline basement systems, the drinking water standard (5µg/L) can naturally be exceeded. High natural background levels are expected nearby mining veins. Antimony is often associated to lead. Examples: - France: In the French Alps, high NBLs of antimony have been observed. For example, a

concentration of 7µg/L has been measured in a well intended to provide drinking water. - Other significant values (around 10-12µg/L) have been measured in the alpine part of

Corsica (northern part of the island).

6.9.2.4. Boron Boron is not very abundant in igneous rocks (table 6.4). The main mineral containing B in these rocks is the tourmaline in which B is substituted to Si. Hence, boron NBLs in crystalline basement aquifers are relatively low. Natural concentrations don’t often exceed a few µg/L.



6.9.2.5. Fluorine Fluorine is relatively abundant in igneous rocks (table 6.4). In this kind of rocks, fluorine is present in fluorspar CaF2, fluorapatite and in amphiboles or micas. But the solubility of these two minerals is very low. Although fluorine is abundant in crystalline basement rocks, its mobilization is quite difficult at low temperature. High NBLs require soluble fluorspar veins occurrences. F- adsorption on some silicates and its release also contribute to high natural concentrations. For all these reasons, fluorine concentrations are very heterogeneous in this kind of aquifers. Examples: - France:

. the previous covering layers of the hercynian basement present high natural background levels of fluorine. It is the case of the Infra-Lias layers in the Morvan massif and of the Infra-Toarcian levels in Poitou area (mean value = 3mg/L),

. in the Mont-Blanc and Aiguilles-Rouges massifs, high natural concentrations have also been measured (up to 4mg/L; Dubois and Parriaux, 1990).

- Belgium (Flanders): the observed fluoride concentrations in the Palaeozoic Basement Aquifer range from 0.05 to 30.9mg/l. Compared to the drinking water limit of 1.5mg/l, the concentrations are mostly exceeding this value.

Figure 6.25 - Distribution of fluoride concentration in the Palaeozoic Basement Aquifer in Flanders.

The distribution of fluoride concentrations is shown in the following figure. Low concentrations are found in the southeastern part of the area (mostly in the province of Flemish-Brabant). Here, the drinking water limit is not exceeded. The low concentrations are found in the CaHCO3-watertype. The elevated calcium concentrations in this zone do not allow fluoride to be high, as this would lead to the precipitation of the mineral fluorite. The high fluoride concentrations in the major part of the area are found in very soft waters, of the NaHCO3- and the NaCl-watertype, in which Na/Ca-exchange has played a major role in determining groundwater quality. The calcium concentrations in these zones are below



10mg/l, allowing for the fluoride concentrations to rise. The majority of fluoride values is around 4mg/l. The origin of the fluoride is natural; it is derived from fluorite and fluoro-apatite minerals in the Palaeozoic rocks. 6.10. Natural background levels in schists and shales aquifers Natural Background Levels can be assumed in this type of aquifer to NBLs of crystalline basement rocks systems. 6.11. Natural background levels in volcanic aquifers Volcanic rocks have a wide range of chemical, mineralogic, structural, and hydraulic properties, due mostly to variations in rock type and the way the rock was ejected and deposited. Unaltered pyroclastic rocks, for example, may have porosity and permeability similar to poorly sorted sediments. Hot pyroclastic material, however, may become welded as it settles, and, thus, be almost impermeable. Silicic lavas tend to be extruded as thick, dense flows and they have low permeability except where they are fractured. Basaltic lavas tend to be fluid and they form thin flows that have considerable pore space at the tops and bottoms of the flows. Numerous basalt flows commonly overlap and the flows are separated by paleosols or alluvial material that forms permeable zones. Columnar joints that develop in the central parts of basalt flows create passages that allow water to move vertically through the basalt. Among the volcanic rocks only the basic volcanites are of supraregional significance and represent the most productive aquifers in volcanic rocks. In Germany, above all, observation wells from the basaltic rocks of the Vogelsberg in Hesse, which is of significance for water management, belong to this unit. The basic volcanites are predominantly basaltoids originating from the Tertiary. In most places, the thickness of these aquifers is 100m or less. Groundwater flow in the basaltic rock aquifers is local to intermediate. In Hesse, the basaltic rock aquifers are quite permeable and represent important aquifers for the regional water resources management. Due to the small number of existing observation wells, no "Acid Volcanites" reference unit was specified for Germany. Although differences according to their mineral composition exist for igneous and metamorphic rocks, the aquifers in these rocks should mainly be classified with regard to their hydrologic regime. The crystalline rock aquifers are in general characterized by significant direct runoff portions, a short residence time of groundwater in the aquifers and thus a low total solution content. In contrast, the volcanic rocks represent, depending on their permeability, regionally significant aquifers, where groundwater recharge plays an important role. This in combination with the presence of more easily weatherable volcanic glass leads to higher TDS in groundwater.



Parameter N

Na mg/l 540

K mg/l 544

Mg mg/l 549

Ca mg/l 554

Fe mg/l 275

Mn mg/l 541

HCO3 mg/l 537

SO4 mg/l 549

Cl mg/l 555

NH4 mg/l 507

NO2 mg/l 656

NO3 mg/l 639

PO4 mg/l 19

DOC mg/l 8

LF µS/cm 546

O2 mg/l 551

H µg/l 561

pH -

F mg/l 329

Br mg/l 17

J mg/l 6

Ag µg/l 87

Al µg/l 540

As µg/l 198

B µg/l 56

Ba µg/l 97

Bi µg/l 16

Cd µg/l 349

Co µg/l 18

Cr µg/l 337

Cu µg/l 36

Hg µg/l 335

Li µg/l 40

Ni µg/l 338

Pb µg/l 348

Sb µg/l 70

Se µg/l 65

Si mg/l 41

Sn µg/l 16

Sr µg/l 40

Zn µg/l 56

Komponentensep.

10. P 50. P 90. P 4,2 6,3 9,3

0,57 0,91 1,4

6,9 12 22

13 27 54

0,01 0,01 0,03

0,004 0,009 0,02

63 138 303

2,5 8,0 26

3,5 8,7 22

0,002 0,006 0,02

0,0001 0,0002 0,0004

4,7 7,8 13

155 273 482

4,8 8,3 10,4

0,02 0,06 0,14

6,8 7,3 7,7

6,1 8,7 12

1,2 6,6 36,1

0,02 0,03 0,04

0,01 0,01 0,02

0,01 0,02 0,03

0,05 0,08 0,13

0,04 0,05 0,07

Figure 6.26 - Natural background concentrations in volcanic rocks in Germany (Kunkel et al., 2004).



6.11.1. Trace elements Volcanic aquifers don’t often present high NBLs. For some element such as B, Ni, or Cr, significant occurrences could appear. But it is important to note that a lack of knowledge exists about trace elements in volcanic aquifers.

6.11.1.1. Boron Boron is sometimes present in volcanic rocks. It is in that case associated to U-Mo-Zn and Ag-Au-Zn groups.

6.11.1.2. Nickel In basalt aquifers with olivine, nickel occurrences can be observed. For this reason, natural occurrences of nickel could occur in volcanic area made of basalts. But it seems like the concentrations remain very low (only a few µg/L). Examples: - France: NBLs around a few µg/L have been measured and are expected in the volcanic

aquifers of Central France. For example, a concentration of 1.1µg/L has been measured in a basaltic massif (Coirons, Ardèche).

- Bulgaria: concentrations between 2 and 2.5µg/L have been observed in volcanic aquifers.

6.11.1.3. Chromium The main mineral containing Cr is chromite FeCr2O4. Cr is easily substituted to Fe and is therefore often present in silicata rich rocks. A volcanic context appears to be in favour for Cr occurrences (Robertson, 1991). It has also been demonstrated that significant Cr concentrations require an oxidized environment and an alkaline pH. A positive correlation with U and Va would also exist (Robertson, 1991). 6.12. Synthesis

6.12.1. Major elements This review of the concentrations that can naturally be found in each type of aquifers has allow to define the range of value than can naturally been expected in each type of aquifers. Of course, these statistics don’t aim to definitively define NBL of European aquifers but for the study of GWB in which no data is available they can be useful. This is besides what is recommended in the methodology developed in WP3 (see report D15 of the project).

6.12.2. Trace elements As a conclusion on trace elements, the following table aims to give an overview on which elements are expected to be present with high NBLs in each type of aquifer. Obviously, as the chemical composition of an aquifer is the result of complex processes (water – rocks interactions, redox conditions, flows, etc.), the goal of this table is not to explain all the factors involved. But it could be useful for water managers in order to quickly identify if a significant concentration of an element could be of natural origin. Of course, as a next step, a more detailed study is necessary to understand the origin of a given element.



Table 6.5 - Relation between aquifer’s types and trace elements occurrence (to be completed).

Limestones Chalk sands and gravels sandstones

oxidized reduced oxidized reduced

Volcanic rocks

Crystalline basement

oxidized reduced oxidized reduced As - - - - - Cd - - - - - - - - Cr - - - - - - - Hg ? ? ? - ? ? ? ? Ni - - - - Pb - - - - - - - - Sb - - ? - - - - Se - - Al - ? - - Ag - - - - - - - - Ba B - - - Cu - - - - - - - F - Fe - ? - - Mn - ? - - P - - - - - - - - Zn - - - - - - - - NH4

: occurrence possible and potentially significant in term of concentrations (although it doesn’t mean that the element will be present systematically).

- : geological context not particularly appropriate for significant presence of the element ? Lack of knowledge Finally, this review of the high concentrations that can naturally be found in European aquifers allows to identify the elements that will have relatively frequently high NBL and that will be important to take into account when defining the good chemical status of groundwater bodies. These elements and the countries concerned by high NBL are summarised in the following table.

Table 6.6 - Trace elements for which high NBL can be expected in European groundwater bodies.

As FR, DK, UK, HU, PL, Belgium-Flanders,… Fe, Mn All countries Ni FR, UK, DK, DE Al LT Sb FR Se FR… F FR, UK, LT, PL, EE Ba EE, LT, FR, DK… B FR, DK, … Al LT, DK, Mo LT NH4 PL, DK…. Cl, SO4 FR, UK, PL, EE



6.13. Saline influence

6.13.1 Salt water Appelo & Postma (1993) mention six processes that may increase the salt content of water to such extent that it is appropriately termed “salt water”. These processes are: - evaporation and concentration / precipitation of dissolved salts; - dissolution of mineral salts; - mixing with recent sea water; - mixing with old water (connate from buried marine sediments); - volcanic exhalations; - hyperfiltration or membrane filtration (reverse osmosis): small pores are completely filled

with a negative potential, such that only water molecules and uncharged species can pass the pore, leaving a concentrated solution behind.

Whereas both latter processes can be considered to be special cases, the four former ones have a wide-spread occurrence, and will be discussed below. Salt waters with a salinity exceeding the one of seawater (TDS 35,000ppm) are called brines. Sedimentary basins may contain brines. Total concentrations as high as 400,000ppm have been reported (Richter & Kreitler, 1993). Though most basins are composed predominantly of marine sediments, formation waters generally do not resemble seawater in either chemical composition or concentration. Two different types of brines are generally found: Na-Cl brines and Na-Ca-Cl brines, neither having chemical composition ratios similar to seawater (Richter & Kreitler, 1993). The origin of the chemical composition of brines in sedimentary basins is widely discussed. Three mechanisms have been used to explain the high ionic concentrations and the chemical composition of the brines (Richter & Kreitler, 1993): the brines originated as a residual brine solution left after the precipitation of evaporates: - 1) basinal waters have dissolved halite; - 2) basinal waters have been forced through low-permeability shales (reverse osmosis); - 3) Land (1987) suggested that the primary source of salinity are evaporite deposits. Furthermore, rock-water interaction is contributing to the composition of brines (Richter & Kreitler, 1993). Mineral equilibria have been invoked for explaining the large dominance of Na and Ca over other cations in almost all brines: equilibrium conditions between Na-feldspar (albite) and K-feldspar result in Na dominance over K, while equilibrium conditions between calcite and dolomite result in Ca dominance over Mg (Helgeson, 1972; Land, 1987). Besides, the massive destruction of detrital feldspar during diagenesis releases significant amounts of Ca, K, Sr, Ba, Li, etc. to the solution. Metamorphism in the basement underlying the sedimentary basin have been proposed to cause large-scale material transport, of components such as SiO2 and Al2O3, and of acidity, into the overlying strata (Land, 1987).

6.13.1.1. Evaporative concentration Evaporative concentration removes water molecules, leaving a concentrated solution behind. The linear and parallel increase of all solutes concentration is interrupted as soon as the solubility product of a mineral is reached, and precipitation of the salt takes place. Upon



continued evaporation, the concentration will continue to increase for solutes that have a concentration ratio in the solution that is higher than the concentration ratio in the precipitating salt. Reversely, for solutes with a lower concentration ratio in the solution compared to the precipitate, the concentration will decrease during further evaporation. Concentration ratios in the solution of the ions composing the precipitating salt, will thus change during evaporative concentration and precipitation (Appelo & Postma, 1993). Hardie & Eugster (1970) have applied the ratio principle to evaporating waters, based on the idea that the composition of salt water and brines is determined by the concentration ratios of the ions in the initial solution. The evolution of the evaporating water composition is shown in figure 6.27. Calcite will be the first mineral to precipitate, and depending on the ratio 2Ca/HCO3 one of both ions will disappear from the solution. The next step will be gypsum (CaSO4.2H2O) precipitation (if 2Ca was dominant over HCO3) or sepiolite (MgSi3O6(OH)2) precipitation. This results in compositions with the remarkable absence of one of the major ions (Appelo & Postma, 1993).

Figure 6.27 - Some possible pathways for evaporation of natural waters according to the model of

Hardie and Eugster (1970) (Appelo & Postma, 1993). From evaporating seawater, calcite is precipitated first, followed by gypsum. In the next step, halite (NaCl) is precipitated, which is (much later) followed by Mg-salts. Only at extreme concentrations, KCl will precipitate. Evaporative concentration takes place in salt flats, lagoons... Salt waters resulting from this process have been formed at or near the ground surface, but may have been moved to greater depths by groundwater flow. Besides in fully natural conditions, this process may also be anthropogenically induced, when irrigation is applied in (semi-)arid areas. An example of natural salinization of groundwater, controlled by evaporative concentration and concomitant gypsum precipitation, resulting in a shift from low-TDS mixed-cation mixed-anion-type water to high-TDS NaSO4 water, is given in Figure 6.28.



Figure 6.28 - Natural salinization of groundwater in parts of the San Joaquin Valley of California,

controlled by (a) evaporation and (b) gypsum saturation, resulting in (c) a shift from low TDS mixed-cation mixed-anion type water to high TDS, NaSO4 water (Richter & Kreitler, 1993 after Deverel &

Gallanthine, 1989)



6.13.1.2. Dissolution of mineral salts As a result of the high solubility of mineral salts, groundwater percolating through or against salt deposits, will readily dissolve them; the major groundwater composition will be largely determined by the ions dissolved from the salt. The largest occurrences of salt deposits consist predominantly of halite, the solution of which produces NaCl-water of relatively uniform chemical composition (Richter & Kreitler, 1993). This is shown by the bivariate plots of major ions and Br/Cl ratios versus chloride for natural halite-solution brines in figure 6.29 (Richter & Kreitler, 1993).

Figure 6.29 - Bivariate plots of major ions and Br/Cl ratios versus chloride for natural halite-solution

brines from Canada, Kansas and Texas. Relatively little scatter in these plots suggests little variation between halite-solution brines from different areas. The composition of laboratory solutions of mined halite deviates from natural solution of salt beds because of the absence of associated evaporates.

(Richter & Kreitler, 1993).



Benavente et al. (1995) describe saline groundwaters in the Antequera Region in Central Andalusia (Southern Spain), resulting from the dissolution of evaporites in Triassic materials. This leads to springs and diffuse outflow of brackish waters of CaSO4-type (TDS between 1.7 and 2.8g/l) and NaCl-brines (TDS between 3 and 90g/l). In the closed depressions resulting from dissolution of the evaporites, salt lakes have developed, in which the climatic conditions with high evaporation rates, have caused the salinity still to increase. Magnesium salts have been reported to be present in the surface crust of the lakes (Benavente et al., 1995).

6.13.1.3. Mixing with recent seawater Coastal aquifers are mostly characterized by the confrontation between marine and continental conditions. This may result in the salinization of fresh aquifers, or conversely, the freshening of saline aquifers (cf. chapter 9.2.1). Salinization can be induced by natural events, such as marine transgressions and flooding, or by anthropogenic causes, such as overexploitation of the aquifer. Freshening can have natural origins, such as the development of dune belts along the coastline, inducing rain-water recharge, or it can be triggered by man by artificial recharge. The mixing ratio of both end members, seawater (NaCl-type) and fresh recharge water (mostly CaHCO3-type, resulting from calcite dissolution in infiltrating rain water), can be assessed from the concentration of the conservative Cl-ion, assuming this parameter is fully delivered by the marine end member. Next to the composition of both end members, also the occupation of exchange sites in the solid matrix is of concern, and will be different for salinization compared to freshening: in the former case, the exchange complex is filled by Ca, while in the latter case, the exchange sites are occupied by marine cations (mainly Na, and also Mg and, to a minor extent, K), reflecting the formerly prevailing conditions. The change towards a new equilibrium will cause cation exchange, as a major reaction to be considered. It is resulting in the typical CaCl-watertype in the case of seawater intrusion in a fresh-water aquifer. The opposite process of freshening produces the typical NaHCO3-watertype, and in a further freshening stage, even the MgHCO3-watertype results (cf. chapter 9.2.1). Calcium ad/desorption, associated with freshening/salinization, may lead to sub/supersaturation with respect to calcite, resulting in calcite dissolution/precipitation. In freshening conditions, this second stage of calcite dissolution may produce very high HCO3-concentrations in the NaHCO3-watertype (Walraevens, 1990). The shifts in relative ion composition can be visualised on piper diagrams (figure 6.30). Cation exchange causes water compositions to deviate from the mixing line between the salt and fresh water end member. It should be emphasized that the role of cation exchange has often been underestimated. Solid carbonate reactions and carbonate dissolution reactions have been reached for to explain Mg-increases that were caused by cation exchange (Appelo, 1994). Along the Belgian coast the upper aquifer system has been salinized in the Holocene by transgression periods. At that time the aquifer became filled with seawater. Later on, dune development in a coastal belt has triggered replacement of the salt water with fresh infiltration water, developing a fresh water bell under the dunes. More inland, in the drained polders, marine waters still prevail. On a hydrochemical cross-section perpendicular to the coastline, typical water types can be recognized (figure 6.31). Under the dunes (left side of profile) fresh CaHCO3 waters can be found till the bottom of the aquifer. Here marine conditions are completely flushed out. At the dune-polder border NaHCO3 waters occur indicating the salt water here is replaced by fresh water, but cation exchange (calcium is replaced by sodium) proves that marine traces still remain in the sediments. Further inland



under the polders, salt NaCl water is still present. Interestingly, the presence of a polder ditch has imposed a local groundwater cycle draining more fresh water from the dune-polder boundary inland and upward to the surface. This lens can be recognized in water samples and in geophysical well loggings. Fresh/salt water distribution is often influenced by local groundwater flow cycles.

Figure 6.30 - Piper diagram showing the shift in ion distribution due to salinization and freshening.

Figure 6.31 - Hydrochemical profile near Wenduine along the Belgian coast (Walraevens et al., 2000)



6.13.1.4. Mixing with old seawater Old seawater may be present in the rocks, with a synsedimentary origin (connate seawater), or resulting from a postsedimentary former marine transgression. Apart from the age of the seawater end member, the processes of groundwater quality change are comparable to the previous case. Figure 6.32 shows measured concentrations of main ions in the Ledo-Paniselian aquifer along a flow line, starting at the recharge area. This aquifer is a typical example of a freshening aquifer, in which fresh recharge water (since Pleistocene times) is gradually replacing the connate Tertiary seawater. Downstream increasing Cl- and SO4

2- concentrations show the growing admixture of the saline end member. Also Na+ exhibits this effect, but additionally, Na+ is increased upstream of the fresh/saline-water interface, as a result of cation exchange. Further upstream, also K+ peaks due to cation exchange, and subsequently, Mg2+ is raised by the same process. Finally, in the most upstream waters, the effect of cation exchange has been reduced, and Ca2+ concentrations are no longer lowered by its effect. Notice the difference in concentration scales: Na+ is more strongly raised by cation exchange than K+ and Mg2+ are. The impact of cation exchange process can also clearly be seen on the Na/Cl ratio (figure 6.33). This ratio is strongly increased in NaHCO3 type waters, originated from CaHCO3 waters by exchange of calcium by sodium.

Figure 6.32 - Changing ion concentrations in the Ledo-Paniselian aquifer along a flow line, with

increasing distance from the recharge area. Crosses represent measured concentrations (Walraevens et al., 2006).



6.13.2. Mixing of salt water with fresh water Salt water, resulting from the processes discussed above, may be brought into contact with fresh water by natural or anthropogenic processes (e.g. pumping), resulting in salinisation or freshening (cf. chapter 9.2.1) as a result of mixing of both end members and rock-water interaction. The presence of a saline end member in groundwater is mostly decisive for its chemical characteristics. This has to do with the concentration difference between fresh and salt water, and the notion about what is understood by “fresh” and “saline” water. A mixture of 95% fresh water (TDS: 375 mg/l) with only 5% seawater (TDS: 35,000 mg/l) produces fairly brackish water with a TDS of 2,106 mg/l, which is certainly no longer recognized as “fresh”. On the other hand, a mixture of 95% seawater with 5% fresh water has a TDS of 33,269mg/l, and is considered salt without any reserve. This involves that freshening has a longer time scale upon which to act, than salinization (Walraevens & Van Camp, 2005). Mixing trends can be evaluated best using the most conservative constituents dissolved in groundwater, that is, chloride and bromide. These constituents are often useful to not only identify the mixing source of salinity, but also to estimate the mixing ratio (Richter & Kreitler, 1993). On the condition that the end members’ composition is known, the composition of the conservative mixture can be calculated, and the extent of reactions can be evaluated comparing the actual composition with the conservative mixture (Appelo & Postma, 1993). Cation exchange is a major reaction to be considered (see above). Calcite dissolution/precipitation may be very important as well. In the case of seawater intrusion, cation exchange causing a Ca-surplus may lead to supersaturation and calcite precipitation, while freshening leads to a Ca-deficit and undersaturation, and a second stage of calcite dissolution will take place, producing very high HCO3-concentrations in the NaHCO3-watertype (Walraevens, 1990). Furthermore, the mere mixing of two groundwaters with different CO2-pressures, both at equilibrium with calcite, will always result in undersaturation and calcite dissolution (Appelo & Postma, 1993). This effect has been called “Mischungskorrosion” in German literature (Bögli, 1978). Mixing of fresh and seawater near the coast might explain the abundant cave formation which is observed in many calcareous coasts. Undersaturation as a result of mixing is common for many minerals (carbonates, silicates and oxides/hydroxides) (Appelo & Postma, 1993). For simple salts, as for gypsum, however, mixing of two saturated waters will always give a supersaturated mixture. Supersaturation will also be found in mixed waters, when other acids besides CO2 play a role, and the CO2-pressure of the more acid water is the lowest (Appelo & Postma, 1993). The question may rise how relevant the chemistry of deep salt waters may be for the more shallow aquifer systems that are of direct concern to the EC-policy. However, deep salt waters have been shown to influence shallow aquifers in natural conditions, even without considering anthropogenic effects, such as pumping and oil exploitation. Fidelibus & Tulipano (1996) mention a hypothesis for the mobility of deep salt waters. In a sedimentary basin, the compaction of argillaceous sediments may be the principal cause for fluid flows which, in the course of geological time, can move large quantities of water. Besides the compaction by simple gravitational loading, a more drastic compression of the sediments can be produced by horizontal compressive stresses of tectonic origin. This would cause the migration of deep fluids which are forced to rise, and may be emerging as springs. Fidelibus & Tulipano (1996) give the example of the Trani and Taranto springs, Apulia, Southern Italy, through which 6*106m3 of old salt waters are outflowing per year, at the Ionian Sea coast.



6.13.3. Tools for detecting sources of salinization It is often possible to use chemical criteria to distinguish between sources of the salt. Richter & Kreitler (1993) list a variety of chemical constituents and constituent ratios that have been used as possible tracers of salinity sources (table 6.7). Parameters most often used include the major cations Ca, Mg, Na, K, the major anions Cl, SO4, HCO3, some minor elements (Br, I, Li) and some isotopes (18O, 2H, 3H, 14C). Some of these constituents are more useful than others (Richter & Kreitler, 1993), and the usefulness also depends from case to case.

Table 6.5 - Geochemical parameters used for identification of salinity sources (Richter & Kreitler, 1993).

6.13.3.1. Major ion chemistry Major ion chemistry is the first information source for deducing the source of salinity. Molar ratios of major chemical constituents, such as Na/Cl, Ca/Cl and Mg/Cl, can be used to differentiate an evaporation trend (1:1 slope) from a mixing trend (typically not a 1:1 slope) (Richter & Kreitler, 1993). Moreover, when evaluating a mixing trend, it should be taken into account that Na, Ca, Mg may have been subjected to cation exchange, such that their ratio to chloride will not just reflect a linear relation between both end members. The Na/Cl-ratio in fresh water and in seawater are basically the same (molar ratio ≅ 0.85); in the case of



freshening, the ratio is greatly increased in the NaHCO3-watertype (Custodio, 1987). This is shown in figure 7 for the Ledo-Paniselian aquifer, as a typical example of a freshening aquifer (Walraevens & Van Camp, 2005).

0.1 1 10 100 1000Cl (meq/l)

0

2

4

6

8

10

Na/

Cl r

atio

Na/Cl ratio versus Cl concentration in the Ledo-Paniselian groundwater

NaCl waters

CaHCO3 waters

MgHCO3 waters

NaHCO3 watersC

ATI

ON

-EXC

HA

NG

E

RECHARGE AREA

MIXING WITH OLD SALINE WATERS

OLD CONATE WATERS

Figure 6.33 - Na/Cl ratio versus Cl concentration in the Ledo-Paniselian groundwater (Walraevens &

Van Camp, 2005) Dissolution of mineral salts may be recognized by typical ratios for major ions. The typical Na/Cl ratio for halite-dissolution brines is 0.64 molar ratio, and for many oil-field/deep-basin brines it is <0.50 molar ratio (Richter & Kreitler, 1993). These molar ratios are quite different from the one for seawater (0.85), allowing to discriminate the source of salinity. Consecutive reactions may however alter the ratio.

6.13.3.2. Bromide Bromide, in combination with chloride, is generally a good tracer of salinization sources since both constituents are relatively conservative (Richter & Kreitler, 1993). In seawater 65 mg/l Br is present (Richter & Kreitler, 1993; Hem, 1985), while in fresh water the concentration of bromide is generally between 0.005 and 0.150 mg/l (Hem, 1985). This difference in contents in seawater and fresh water makes bromide a good indicator to identify intrusion. High bromide concentrations can also occur in geothermal waters (several tens of mg/l) and in brines (several tens to more than 2,000mg/l) (Richter & Kreitler, 1993). The Br/Cl ratio for seawater and most fresh waters is 3.3*10-3 weight ratio (Richter & Kreitler, 1993), or 15 * 10-4 – 17 * 10-4 molar ratio (Custodio, 1976). In order to avoid using such small numbers, Alcalà & Custodio (2005) mention the reciprocal number, Cl/Br molar ratio for seawater at 655 ± 4. Halite-dissolution brines have a Br/Cl ratio which is one order of magnitude smaller than in seawater: <5 * 10-4 weight ratio (for example see figure 3). This is due to the fact that only small amounts of Br are incorporated into the crystal structure of halite during evaporation.



Oil-field/deep-basin brines, on the other hand, have a Br/Cl ratio which often is significantly higher than seawater, and is typically one or more orders of magnitude higher than in halite-dissolution brines (Richter & Kreitler, 1993). Pollution, due to addition of wastewater or leaching of wastes, produces generally a clear decrease in Br/Cl up to <2.25 * 10-3 weight ratio (Alcalà & Custodio, 2005). Pollution by road-deicing salts brings Br/Cl even down to <1 * 10-3 weight ratio (Richter & Kreitler, 1993). The former use of obsolete (now forbidden) Br-containing pesticides (methyl bromide type, used against nematodes) causes increased Br/Cl (Alcalà & Custodio, 2005). Differentiation of salt-water sources using Br/Cl ratios works best at high concentrations of TDS. Morell et al. (1986) used Br/Cl to determine the source of salinity in the coastal area of Torreblanca, and found that both a marine component and a component from deep brines were present. Ivask et al. (2003) found the Cambrian-Vendian aquifer in Estonia to have high Br/Cl ratios compared to seawater of the Baltic Sea, which indicates an additional bromide source is present in these sediments. They found the bromide enrichment in groundwater relative to seawater is probably due to ascending, saline bedrock groundwater, rather than due to lateral seawater intrusions.

6.13.3.3. Strontium and 87Sr/86Sr Strontium is soluble in aqueous solution as the +2 ion and is geochemically very similar to Ca2+. In fresh water strontium is only present in small concentrations due to the scarce presence and low solubility of minerals containing strontium as the major cation, like celestite (SrSO4) and strontianite (SrCO3). Seawater has strontium concentrations in the order of 9-14 mg/l (Morell et al., 1986). Fidelibus & Tulipano report 7-8 ppm for seawater. Strontium is involved in processes like ion-exchange and chemical precipitation. It will be removed from solution, when encountering water containing sulphates. According to Fidelibus & Tulipano (1996), the Sr/Cl ratio in subsurface waters of ancient basins is always higher than that of modern seawater: next to simple dissolution of carbonates, incongruent dissolution of Sr-rich skeletal aragonite raises the Sr-concentration in solution at each dissolution-precipitation cycle (Plummer et al., 1976). The Sr (and also Ba) contained in brines of sedimentary basins are known to be vastly more abundant than predicted simpy by the evaporation of seawater, and the 87Sr/86Sr ratio of the brine is commonly elevated (Richter & Kreitler, 1993). This is attributed by Land (1987) to rock-water interaction, namely the significant involvement of silicate phases in determining brine chemistry. The massive destruction of detrital feldspar releases significant amounts of calcium, potassium, strontium, barium, lithium, etc. into the solution. 87Rb, which is a characteristic element of silicate phases, especially K-feldspar, is producing 87Sr by decay. From the four (stable) Sr-isotopes, only 87Sr can vary with respect to the other isotopes, due to β-decay of 87Rb, but the variations are small due to both the low Rb abundance and the very long 87Rb half-life (48.8 billion year). The 87Sr/86Sr ratios of dissolved Sr in groundwater are mainly controlled by water-rock interactions and, therefore, the isotopic compositions of



the strontium bearing minerals in the host rock (Faure, 1986). Important Sr-bearing minerals are carbonates (calcite, aragonite and dolomite) and evaporates (gypsum and anhydrite), and also silicates (feldspars, hornblendes, micas and clays) (McNutt, 2000) to a lower extent. Sr can also enter into the lattice of fluorite and barite. A mineral has a given 87Sr/86Sr value at the time of its formation and that ratio will increase over time in response to 87Rb decay (McNutt, 2000). It is known that Rb constitutes for K, and Sr for Ca in minerals. Limestones and dolomite contain little Rb and the bulk of the carbonate rocks has a marine origin meaning that the 87Sr/86Sr ratio is always low, less than 0.710 for unaltered rocks (McNutt, 2000). Carbonate minerals are more soluble than most silicates and enriched in Sr. Groundwater flowing through carbonate aquifers will reach high concentration levels of Sr, resulting in fluids with low 87Sr/86Sr. This is in contrast to silicate groundwaters, which tend to have higher 87Sr/86Sr and lower Sr concentrations. Starting from the very low value of recharge water, Sr-concentrations in groundwater in carbonate aquifers increase due to rock-water interaction (Barbieri et al., 1998). Also the Sr-isotopic composition of the minerals will tend to be reflected in the water. Consequently, natural waters in carbonate aquifers exhibit a wide range of both the Sr-concentration and the Sr-isotope ratio, according to the different lithological characteristics, the residence time, and the age of the rocks met during their underground evolution. The longer the residence time of groundwater, the closer the chemical equilibrium with the solid matrix will be. Residence time being equal, Sr-enrichment depends on the mineralogical characteristics of the rocks. For similar mineralogical characteristics, different ages of the rock will result in different Sr-isotope ratios (Barbieri et al., 1998). The87Sr/86Sr-ratio methodology has been applied in a variety of environments to study surface waters and groundwaters, including brines (Barbieri et al., 1998). Main references are in Banner (1995) and in Banner et al. (1989). Blum & Chery (see annex) established a steady increase in Sr along the flowpath in the Jurassic limestone aquifer in the Aquitaine Basin, France. The 87Sr/86Sr ratio was interpreted to be pointing to mixing with water from the (probably) marly Triassic at one hand, and from the carbonated Triassic on the other hand (figure 6.34). Shand et al. (2003) explained the increase in salinity in the Triassic sandstone of the Vale of York, close to the confined margin with the Mercia Mudstone, by the dissolution of evaporates at the base of the confining Mercia Mudstones Group, based on high Sr concentration combined with the dominance of Ca and SO4 and high SO4/Cl ratios.



Figure 6.34 - Sr-isotope ratio vs. 1/Sr in the Jurassic limestone aquifer in Aquitaine Basin, France

(Blum & Chery, see annex).

6.13.3.4. Lithium Li is not incorporated into the crystal lattice of minerals, except for clay minerals, which selectively remove Li from water, while themselves becoming Li-enriched; however Li is unstable in the lattice structure due to its charge. A small rise in temperature is enough to cause Li to leave the mineral: a very small difference in temperature at which the rock-water interaction takes place, brings about an increase of the Li-content in water of a factor of 100 to 10,000 times. Contents in thermal waters are in fact extremely high: 1,000 ppb are normal for thermal springs, reaching up to 10 ppm (Fidelibus & Tulipano, 1996). Besides temperature, also squeezing through clay membranes produces squeezed fluids selectively enriched in Li: Li passes through the membrane preferentially, with respect to other monovalent and divalent cations (Kharaka & Berry, 1973). Once lithium is brought into solution by weathering reactions, it tends to remain in the dissolved state since common ion-exchange minerals apparently adsorb lithium less strongly than they do other elements (Hem, 1985). Lithium will thus accumulate in solution depending on time and availability in the host rock (Richter & Kreitler, 1993). Therefore, it may be used as a good indicator of the degree of water-rock interaction. Lithium is found in evaporates and natural brines (Hem, 1985). In fresh and brackish groundwater, Li is in the µg/l range. This indicates that Li is best used at high salt concentrations. Whittemore & Pollack (1979) observed a large range of Li/Cl ratios in oil-field brines as a whole, but a narrow range in a particular geographic region. The Li/Cl ratio may thus be a good indicator for local salinization sources (Richter & Kreitler, 1993).



In modern seawater, the Li/Cl weight ratio is very low, ≅ 10-6, while in many oil-field waters, this ratio increases up to 10-3 (Fidelibus & Tulipano, 1996). In his study of salt and brackish groundwaters of northwestern Germany, Hahn (1972) discusses the Li-Cl relation. He reports Li-concentrations of 5-15ppb for fresh water with 20ppm Cl, while seawater with 19,345ppm Cl has 190ppb. For northwestern Germany, the Li-Cl relation allowed him to differentiate mixtures of seawater and fresh water from mixtures between deep salt groundwater and fresh water, and from mixtures comprising all three components (figure 6.35). The Li/Cl-ratio is decreased with respect to seawater in recently intruded seawaters (reaching depths of 250 m in NW Germany), due to uptake of Li in the crystal lattice of newly formed clay minerals. In contrast, deep groundwaters (from >250m depth in NW Germany) show a clear increase in Li, up to 100,000 ppb, leading to a Li/Cl-ratio far above seawater (Hahn, 1972).

Figure 6.35 - Differentiation of salt water with different salinization cause based on lithium and chloride

content (Hahn, 1972) Fidelibus & Tulipano (1990) have observed increasing concentrations of Ca, Sr and Li with increasing salt contents, in spring waters from Apulia (Southern Italy) (figure 6.36). They describe the average fresh-water component with TDS = 500mg/l, Ca = 100mg/l, Sr = 0.5mg/l and Li practically absent. Based on the very high Li/TDS ratio in some springs, a very old salt water component was inferred by these authors. In their continued investigation, Fidelibus & Tulipano (1996) mention three different groups of salt groundwaters in the Apulian region: - seawater of recent intrusion; - salt waters of ancient intrusion, having been subjected to diagenesis essentially in a

carbonate environment: mainly Ca, Mg and Sr are modified with respect to seawater, due to water-rock interaction, resulting in a decrease of the Mg/Ca ration (due to dolomitisation) and Sr-enrichment, in connection with increasing residence times;



Figure 6.36 - Cross plots of Ca2+, Mg2+, Na+, K+, SO4

2-, HCO3- concentrations versus Cl- contents for

the coastal spring waters, brackish and salt groundwaters, and sea water. Lines stand for fresh water – seawater conservative mixing (Fidelibus & Tulipano, 1996)



- very old salt waters, having variable chemical characteristics, and having been subjected to larger diagenetic changes of chemical composition with respect to seawater: all ions are involved and, besides the interaction with carbonate rocks, interaction with clay sediments is needed to explain part of the observed chemical changes. These salt waters of the basement show a variable, but very high, enrichment in lithium.

Most of the coastal springs discharging in the area contain water of the old and very old salt end members (Barbieri et al., 1998).

6.13.3.5. Stable isotopes of the water molecule (18O, 2H) Average seawater, as the standard for expressing 18O and 2H, has by definition δ18O- and δ2H-values equal to 0 ‰. Fractionation during evaporation of seawater leads to water vapour depleted in heavy isotopes compared to seawater (δ18O = -10‰; and δ2H = -80‰). Condensation to rainwater again favours the heavy isotopes, depending on temperature. For coastal seawater, 18O and 2H will show a characteristic relation, while their absolute values will be temperature dependent. This leads to the mean water line (MWL). Oxygen-18 and deuterium can be useful to distinguish local meteoric water that dissolved halite in the shallow subsurface from a regional oil-field brine or any other brine that is derived from a different recharge area. At high temperature, waters will equilibrate with 18O of the aquifer material and therefore will not preserve the original signature. This oxygen shift can be of value in differentiating deep basinal waters from shallow meteoric waters (Richter & Kreitler, 1993). Evaporation can be recognized on the δ18O-δ 2H diagram by deviations from the meteoric water line δ 2H = 8.13 δ18O + 10.8. Evaporation under non-equilibrium conditions will results in a smaller slope, depending on humidity: for zero humidity, the slope is 3.9, whereas for humidity close to 75 %, it is greater than 5; not until humidity reaches 90%, does the slope approach 8 (Clark & Fritz, 1997). An evaporative trend line was inferred from the groundwater samples from San Joaquin valley, shown in figure 2 (Richter & Kreitler, 1993). 18O and 2H behave as conservative parameters, just as Cl. The relation between 18O and Cl may provide interesting information. Mixing of fresh recharge water with seawater results in a mixing line on the δ18O-Cl diagram. Manzano et al. (2002) in their study of the Inca-Sa Pobla Aquifer in Mallorca, deduced the presence of two different fresh recharge waters, after correcting δ18O for saline contribution (figure 6.37). Devos (1984) studied the groundwaters in the coastal dune area of De Haan (Belgium). Brackish waters were found to be mixtures of rainwater recharge and seawater, as deduced from their δ18O-Cl relation (figure 6.38). Four samples taken from near the Tertiary clay substratum pointed to a different saline end member with much higher chlorinity, that could correspond with Tertiary seawater.



Figure 6.37 - δ18O signature of groundwater and of the fresh water component after correcting for

saline contribution (Manzano et al., 2002)

Figure 6.38 - δ18O-Cl in groundwater of the coastal dune area of De Haan (Devos, 1984).

6.13.3.6. Stable isotopes of dissolved sulphates (34S, 18O) Sulphur isotopes may be used to identify the source of sulphate in groundwater, which in turn may be useful to differentiate between salinization sources (Richter & Kreitler, 1993). Sulphur has valence states from +6 to –2 and the higher valence states tend to be enriched in heavy isotopes (Krouse & Mayer, 2000). δ34S values in natural sulphur compounds cover a range from lower than -50‰ to higher than +50‰. The oxygen isotopic composition of sulphate is influenced by the δ18O value of atmospheric O2 and that of water in which it is formed.



Atmospheric O2 has a δ18O value of +23.5‰. The δ18O value of groundwater is usually similar to that of the mean annual precipitation in the recharge area. Evaporite minerals, like gypsum and anhydrite, are the most soluble sulphur compounds. The δ34S value of sulphate in evaporates ranges from +35 to +10‰ (Krouse & Mayer, 2000). Dissolution of evaporates gives the water a characteristic isotopic composition typical for the unit in which the minerals occurs (Richter & Kreitler, 1993). Modern oceanic sulphate has δ34S values near +21‰ (Rees et al., 1978) and a mean δ18O value of +10‰. Sulphate with a marine origin will have the same isotopic signature as seawater. This makes it possible to delineate the saline-freshwater mixing. Sulphate reduction causes an enrichment in 34S in the residual sulphate. Pauwels (see annex) has recognised a paleo-saline (marine) groundwater (Tertiary transgression) as an end member in the deep more saline groundwater of the hard rock aquifer in Brittany, France, based on δ34S values of SO4 amounting up to +20 and even to +22.3‰, in groundwater which has not undergone sulphate reduction. Shallow waters also show increased Cl and SO4 concentrations, but these are due to agricultural contamination; δ34S in these anthropogenically influenced groundwaters is +8 to +15‰, where sulphates are mainly derived from fertilization and/or atmospheric deposition. In case of denitrification, sulphates derived from oxidation of reduced sulphur compounds involves a decrease of δ34S up to –0.10 ‰. Pauwels (see annex) thus concludes that δ34S appears to be a relevant tool to differentiate between the deeper pristine groundwaters, while their Cl/SO4 ratio is comparable. Blum & Chery (see annex), in their study on the Jurassic limestone aquifer from the Aquitaine Basin in France, distinguished sulphates with an atmospheric origin from sulphates resulting from dissolution of Triassic evaporites, based on δ34S and δ18O from SO4 (figure 6.39).

Figure 6.39 - δ34S versus δ18O of sulphates, and sulphate concentration versus δ34S in the Jurassic limestone aquifer in Aquitaine Basin, France (Blum & Chery, see annex).

6.13.3.7. Boron isotope 11B

-10 0 10 20 30

δ34S(SO4) ‰

0

100

200

300

400

500

SO

4 (m

g/L)

CREGLABLACHAPROBOU TOU

ANTLAFPRA

LALBRU

PAS

CAS

REA

reduction

Unconfined

Confined

0 4 8 12 16 20δ18O(SO4) ‰

-10

0

10

20

30

δ34S

(SO

4) ‰

CRE

GLA

BLA

CHAPRO

BOU

TOU ANTLAFPRA

LALBRUPAS

CAS

REA

Sulfates with an atmospheric origin

Sulphates linked with Triassic evaporites leaching

reduction



A large isotopic variation in nature makes boron isotopes useful for tracing the origin of dissolved salts in groundwater. Seawater has high δ11B values of 39‰, while fresh water with a rock-derived source of boron has relative low δ11B values (e.g. lakes in the Alps = 0.9‰ to 6.2‰) (Vengosh & Spivack, 2000). The large difference between marine and non-marine sources allows direct discrimination of these sources in groundwater systems. Vengosh et al. (1994) have shown that adsorption onto clay minerals during the migration of fluids through the vadose zone of calcareous sandstone modifies the original composition of the solution towards higher δ11B values and lower B/Cl ratios. In the saturated zone the δ11B values of fresh groundwater are derived from the host aquifer rocks (usually with low δ11B values) and of boron supplied by marine sources with high δ11B values. The large isotopic variability of the source rocks (marine carbonate: ca. 10 to 20‰; adsorbed boron: ca. 15‰, bulk boron in clay minerals: ca. -20 to 0‰, igneous rocks: -3 to 3‰) enables evaluation of the sources of dissolved boron in the water and the relative contribution of the different rocks (Vengosh & Spivack, 2000).

6.13.3.8. 3H and 14C Age-dating techniques using radioactive isotopes such as 14C and tritium (3H) allow differentiation between older and modern seawater intrusion (Richter & Kreitler, 1993). 14C and 3H are produced by cosmic rays interacting with nitrogen in the outer atmosphere. Because this constant creation of 14C, the ratio 14C/12C in the atmosphere is relatively constant (pre-1950). Groundwater in aquifers below the water table has no contact with the atmosphere and this ratio will decrease as a result of radioactive decay of the 14C isotope. The 14C/12C ratio allows determination of the time span that has elapsed since the isolation from the atmosphere. The internationally accepted half-life of 14C is 5730 years. Carbon isotopes reflect water-rock interactions and thereby may mask the origin of the water (Kalin, 2000). The addition of “dead carbon” form dissolution of carbonate rocks complicates the age determination. Tritium (3H) has a half-life of 12.43 years (Unterweger et al., 1980), which restricts age determination to only a few tens of years. Large amounts of tritium were introduced into the atmosphere by thermonuclear testing in the fifties and sixties, which now enables differentiation of pre-1950 waters (0-2 TU) from post-1950 waters (> 2-3 TU). Because 3H forms part of the water molecule and is geochemically conservative, interpretation is more straightforward than for other dissolved isotopic or chemical tracers (Solomon & Cook, 2000). Tritium can be used to distinguish between recent salinization sources, with measurable tritium activity, and old salinity sources, with no measurable 3H. In northern Estonia the glacial origin of groundwater from the Cambrian Vendian Aquifer in northern Estonia gives a unique isotopic composition and can be used as an indicator for recent marine intrusions. Based on δ18O, 14C and 3H Ivask et al. (2003) found no intrusion of modern seawater as a consequence of heavy pumping into the Cambrian-Vendian aquifer. Instead the elevated salinity is caused by a deeper old source. In his study of the coastal dune area of De Haan (Belgium), Devos (1984) inferred two salt water intrusion stages, considering 14C of groundwater: one intrusion in Subboreal times (corrected ages around 3,200 BP), and one during the Dunkirk II transgression phase (around 1,600 BP).



6.14. References Alcalá, F.J. & Custodio E. (2005). Use of the Cl/Br ratio as a tracer to identify the origin of salinity in some coastal aquifers of Spain. Proceedings of 18th Salt Water Intrusion Meeting, 2004, Cartagena, Spain, 481-497. Appelo, C.A.J. (1994). Cation and proton exchange, pH variations, and carbonate reactions in a freshening aquifer. Water Resour. Res. 30(10), 2793-2805. Appelo, C.A.J. & Postma, D. (1993). Geochemistry, groundwater and pollution. Balkema, Rotterdam, The Netherlands, p. 536. Banner, J.L. (1995). Application of the trace element and isotope geochemistry of strontium to studies of carbonate diagenesis. Sedimentology 42, 805-824. Banner, J.L., Wassenburg, G.J., Dobson, P.F., Carpenter, A.B. & Moore C.H. (1989). Isotopic and trace element constraints on the origin and evolution of saline groundwaters from central Missouri. Geochim. Cosmochim. Acta 53, 383-398. Barbieri, M., Barbieri, M., Fidelibus, M.D., Morotti, M., Sappa,G. & Tulipano, L. (1998). First Results of the Application of the Isotopic Ratio 87Sr/86Sr in the Characterization of Sea-Water Intrusion in the Coastal Karstic Aquifer of Murgia (Southern Italy). Proceedings 15th Salt Water Intrusion Meeting, Ghent. Natuurwet. Tijdschr. 79, 132-139. Benavente, J., Almécija, C. & Carrasco, F. (1995). Origin and Environmental Significance of Saline Waters in the Antequera Region (Southern Spain). In: Cruz-SanJulian, J. J. & Benavente, J. (1995) eds.: Wetlands: a Multiapproach Perspective, Hydrological and Ecological Studies applied to Wetlands Management in Semiarid Climate. T. G. Arte, Granada, Spain, 55-68. Blum, A., Chery, L., Barbier, J., Baudry, D., Petelet-Giraud, E. (2002). Contribution à la caractérisation des états de référence géochimique des eaux souterraines. Outils et méthodologie. Rapport final, Rapport BRGM RP-51549-FR, 5 vol. Blum, A., Chery, L., Petelet-Giraud,E., Negrel, P. (2005). Characterisation of the geochemical baseline of French groundwaters: Isotopic constraints (Sr, N, S, O). Application to the large Jurassic aquifer (SW France). In EGU Conference, Vienna, Austria - 24-29/04/2005. Bögli, A. (1978). Karsthydrographie und physische Speläologie. Springer, Berlin, 292 p. BRGM (2006). Qualité naturelle des eaux souterraines. Méthode de caractérisation des états de référence des aquifères français. Collection scientifique et technique. Ed. BRGM, 238 p + 1 CD-ROM. Clark, I.D. & Fritz, P. (1997). Environmental Isotopes in Hydrogeology. Lewis Publishers, Inc., CRC Press, Florida, 328 p. Coetsiers, M., Van Camp, M., Walraevens, K. (2005). Influence of the former marine conditions on groundwater quality in the Neogene phreatic aquifer, Flanders. In: Araguas L, Custodio E., Manzano M. (eds.) Groundwater and saline intrusion, Selected papers from the 18th Salt Water Intrusion Meeting, Cartagena 2004, 499-509.



Custodio, E. (1976). Hidrogeoquimica. Sect. 10 of Hidrologia Subterranea. Omega, Barcelona, pp. 614-1005. Custodio, E. (1987). Hydrogeochemistry and tracers. In: Custudio, E. & Bruggeman, G. A. (eds.) Groundwater problems in coastal areas. Studies and reports in Hydrology 45, 213-269. Deverel, S.J. & Gallanthine, S.K. (1989). Relation of salinity and selenium in shallow ground water to hydrologic and geochemical processes, western San Joaquin Valley, California. Journal of Hydrology 109, 125-149. Devos, J. (1984). Hydrogeologie van het duingebied ten oosten van De Haan. PhD-dissertation. Ghent University. Edmunds, W.M., Shand, P. (coordinators) (2003). Natural Baseline Quality in European Aquifers: A Basis for Aquifer Management. EC-project EVK1-CT1999-0006. Final Report. British Geological Survey, Wallingford. Faure, G. (1986). Stable isotope geochemistry, 2nd ed. Springer, Berlin Heidelberg New York Fidelibus, M. & Tulipano, L. (1990). Major and Minor Ions as natural tracers in mixing phenomena n coastal carbonate aquifers of Apulia. Proceedings 11th Salt Water Intrusion Meeting, Gdansk, Poland, pp. 283-293. Fidelibus, M.D. & Tulipano, L. (1996). Regional flow of intruding sea water in the carbonate aquifers of Apulia (Southern Italy). Proceedings of 14th Salt Water Intrusion Meeting, Malmö, Sweden, pp. 230-240. Hahn, J. (1972). Diagenetisch bedingte Veränderungen im Chemismus intrudierter Meerwässer und ihre Beziehungen zum Chemismus von Tiefengroundwässern in Nordwestdeutschland. Geol. Jb. 90, 245-264. Hardie, L.A. & Eugster, H.P. (1970). The evolution of closed-basin brines. Miner. Soc. Am. Spec. Publ. 3, 273-290. Helgeson, H.C. (1972). Chemical interaction of feldspars and aqueous solutions. In: Mackenzie, W.S. & Zussman, J. (eds.) Feldspars. Proceedings. NATO Advances Study Institute, Manchester University Press, 187-217. Hem, J.D. (1985). Study and interpretation of the chemical characteristics of natural water, U.S. Geological Survey Water Supply, paper 2254, 264 p. Hinsby, K., Rasmussen, E.S and Henriksen, H.J. (2003a). European reference aquifers: The Miocene sand aquifers of western Denmark. Report for the EU research project (“BASELINE”): “Natural Baseline Quality of European Groundwaters: A basis for Aquifer Management”. Hinsby, K., Troldborg, L. and Purtschert, R. (2003b). European reference aquifers: The Pleistocene sands around Odense, Denmark Report for the EU research project (“BASELINE”): “Natural Baseline Quality of European Groundwaters: A basis for Aquifer Management”.



Hinsby, K., Jensen, T.F. and Bidstrup, T. (2003c). European reference aquifers: The limestone aquifers around Copenhagen, Denmark. Report for the EU research project (“BASELINE”): “Natural Baseline Quality of European Groundwaters: A basis for Aquifer Management”. Hinsby, K., Edmunds, W.M., Loosli, H.H., Manzano, M., Melo, M.T.C. & Barbecot, F. (2001a). The modern water interface: recognition, protection and development - Advance of modern waters in European coastal aquifer systems. In: Edmunds and Milne (Eds.): Palaeowaters in Coastal Europe: evolution of groundwater since the late Pleistocene. Geol. Soc. Spec. Publ., 189, 271-288. Hinsby, K., Harrrar, W.G., Nyegaard, P., Konradi, P., Rasmussen, E.S., Bidstrup, T., Gregersen, U. & Boaretto, E. (2001b). The Ribe Formation in western Denmark: Holocene and Pleistocene groundwaters in a coastal Miocene sand aquifer. In: Edmunds and Milne (Eds.): Palaeowaters in Coastal Europe: evolution of groundwater since the late Pleistocene. Geol. Soc. Spec. Publ., 189, p. 29-48. Ivask, J., Kaup, E., Marandi, A., Martma, T., Randla, V., Vaikmäe, R., Vallner, L. (2003). The Cambrian-Vendian aquifer, Estonia. In: Edmunds W.M. & Shand P. (Eds.) Natural Baseline Quality in European Aquifers: A Basis for Aquifer Management. EC-project EVK1-CT-1999-00006. Final Report. British Geological Survey, Wallingford. Kalin, R.M. (2000). Radiocarbon dating of groundwater systems. In: Cook, P. & Herczeg, A. L. (eds.) Environmental tracers in subsurface hydrology. Kluwer Academic Publishers, 111- 144. Kharaka, Y.K. & Berry, F.A.F. (1973). Simultaneous flow of water and solutes through geological membranes – I. Experimental investigation. Geoch. et Cosmoch. Acta, 37, 2577-2603. Krouse H. R. & Mayer B. (2000). Sulphur and Oxygen Isotopes in Sulphate. In: Cook, P. & Herczeg, A. L. (2000). Environmental tracers in subsurface hydrology. Kluwer Academic Publishers, pp. 195-231. Kunkel, R. Voigt, H.J., Wendland, F., Hannappel, S. (2004). Die natürliche, ubiquitär überprägte Grundwasserbeschaffenheit in Deutschland. Schriften des Forschungszentrums Jülich, Reihe Umwelt / Environment, Band 47, Forschungszentrum Jülich GmbH, Jülich, Germany, 204 S. Land, L. (1987). The major ion chemistry of saline brines in sedimentary basins. In: Physics and Chemistry of Porous Media II (J. R. Banavar, J. Koplik & W. Winkler, eds.) Ridgefield, Conn., American Institute of Physics Conference Proceedings 154, 160-179. Manzano, M., Custodio, E., Riera, X., González, C., Barón, A. & Delgado, F. (2002). Saline groundwater in the Inco-Sa Pobla Aquifer, SE of Mallorca Island (Balearicj Islands, Spain). Proceedings of 17th Salt Water Intrusion Meeting, Delft, The Netherlands, 250-261. McNutt, R.H. (2000). Strontium Isotopes. In: Cook, P. & Herczeg, A. L. (2000). Environmental tracers in subsurface hydrology. Kluwer Academic Publishers, 233-260. Morell, I., Medina, J., Pulido, A. & Fernandez-Rubio, R. (1986). The use of bromide and strontium ions as indicators of marine intrusion in the aquifer of Oropesa –Torreblanca (Castellon, Spain). Proceedings 9th Swim, Delft 13-16 May, 629-640.



Plummer, L. N., Vacher, H.L., Mackenzie, F.T., Bricker, O.P. & Land, L.S. (1976). Hydrogeochemistry of Bermuda: A case history of ground-water diagenesis of biocalcarenites. Geol. Soc. Am. Bull. 87, 1301-1316. Rees, C.E., Jenkins, W.J. & Monster, J. (1978). The sulphur isotopic composition of ocean water sulphate. Geochim. Cosmochim. Acta 42, 377-381. Richter, B.C. & Kreitler, C.W. (1993). Geochemical Techniques for Identifying Sources of Ground-Water Salinization. Smoley, CRC Press, Florida. RIVM (2001). Background concentrations of 17 trace metals in groundwater in the Netherlands. RIVM, report no. 71170101. RIVM (2003). Basis values for elements in fresh groundwater in the Netherlands; data from the national and provincial groundwater quality monitoring networks. RIVM, report no. 714801028. Robertson, F.N. (1991). Geochemistry of ground water in alluvial basins of Arizona and adjacent parts of Nevada, New Mexico, and California, U.S. geological survey professional paper 1406-C, 90 p. Shand, P., Tyler-Whittle, R., Morton, M., Simpson, E, Lawrence, A.R., Pacey, J. & Hargreaves, R. (2003). Baseline Report Series: 1. The Triassic Sandstones of the Vale of York. British Geological Survey, Commissioned Report CR/02/102N. Solomon, D.K. & Cook, P.G. (2000). 3H and 3He. In: Cook P. & Herczeg A. L. (eds.) Environmental tracers in subsurface hydrology. Kluwer Academic Publishers, 397-424. Soveri, J., Mäkinen, R., Peltonen, K, (2001) – Pohjaveden korkeuden ja laadun vaihteluista Suomessa 1975-1999. Changes in groundwater levels and quality in Finland 1975-1999. Suomen ympäristö, Ympäristönsuojelu nro 420, Helsinki ISBN 952-11-0746-4, ISSN 1238-7312. In Finnish with abstract and summary in English. TNO (2004). Feasibility study on region-specific target values for groundwater in Northern and Southern Holland, the Netherlands. TNO Netherlands Institute of Applied Geoscience, report no. NITG 04-184-B. TNO (2005). The concentration of phosphate in regional exfiltration water in the Netherlands. TNO Netherlands Institute of Applied Geoscience, in press. Travi, Y. (1993), Hydrogéologie et hydrochimie des aquifères du Sénégal. Hydrochimie du fluor dans les eaux souterraines, Sciences Géologiques, mémoire n°95, 155 p., Unterweger, M.P., Coursey, B.M., Schima, F.J. & Mann, W.B. (1980). Preparation and calibration of the 1978 national bureau of standards tritiated-water standards. Int. J. Appl. Radiat. Isot. 31, 611-614. Vallée, K. (1999), Le nickel dans les eaux alimentaires. Application à des champs captants du bassin Artois-Picardie, Thèse de doctorat, EUDIL, Université Lille I, 268p.



Vengosh, A. & Spivack, A. (2000). Boron Isotopes in Groundwater. In: Cook P. & Herczeg A. L. (2000). Environmental tracers in subsurface hydrology. Kluwer Academic Publishers, pp. 479-485. Vengosh, A., Heumann, K.G., Juraske, S. & Kasher, R. (1994). Boron isotope application for tracing sources of contamination in groundwater. Environ. Sci. Technol. 28, 1968-1974. Walraevens, K. (1990). Hydrogeology and hydrochemistry of the Ledo-Paniseliaan semi-confined aquifer in East- and West-Flanders. Academia Analecta, 52, 12-66. Walraevens, K., Cardenal-Escarcena, J. & Van Camp, M. (2006). Reaction transport modeling of a freshening aquifer (Tertiary Ledo-Paniselian aquifer, Flanders-Belgium). Applied Geochemistry (accepted). Walraevens, K., Clays S., Hermans, A. & Van Camp, M. (2000). Hydrogeological and hydrogeochemical investigation of the dune area and adjacent low polders at Wenduine-Uitkerke, Flemish Coastal Plain. In Sadurski (ed.) Hydrogeology of the coastal aquifers, Proceedings of the 16th Salt Water Intrusion Meeting, Poland. pp. 161-167. Walraevens K. & Van Camp M. (2005). Advances in understanding natural groundwater quality controls in coastal aquifers. In: Araguàs Custodio & Manzano (eds.) Groundwater and saline intrusion, selected papers from the 18th Salt Water Intrusion Meeting, Cartagena 2004, 449-463. Wendland, F., Hannappel, S., Kunkel, R., Schenk, R., Voigt, H.J., Wolter, R. (2005). A procedure to define natural groundwater conditions of groundwater bodies in Germany. Water Science and Technology 51 (3-4), 249-257. Whittemore, D.O. & Pollack, L.M. (1979). Geochemical identification of salinity sources. In: French, R.H. (ed.) Salinity in watercourses and reservoirs. Proceedings of the 1983 International Symposium on State-of-the-Art Control of Salinity, Salt Lake City, Utah, 505-514.


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