Artificial recharge of aquifers can be performed for various purposes and under varying hydrogeological conditions. We present an overview of deep-well recharge applications which have taken place in the Netherlands over the last two decades. We present the purpose of each application, the issues which had to be resolved, the preventive measures which were taken to improve performance and the lessons learned from each experience. Examples are given of applications which aimed at the storage of water for drinking and other purposes such as irrigation, achieving environmental goals and disposal of wastewater. Applications aiming at drinking water production usually faced issues related to the quality of the abstracted water not meeting drinking water standards with respect to various elements, such as iron, manganese and arsenic. Storage of water in brackish aquifers was complicated by buoyancy effects making part of the recharged water irrecoverable. Recharge of water with the purpose of recovering declined groundwater tables and fighting seawater intrusion was hindered by clogging of the injection well while the disposal of wastewater was limited to aquifers of lower groundwater quality.

INTRODUCTION

Artificial recharge of aquifer systems may be performed for various reasons. These include, among others, the subsurface storage of water for drinking and agricultural purposes, the natural subsurface water treatment, the achievement of environmental goals (e.g. recovery of declining water tables and the prevention of saltwater intrusion), and the disposal of wastewater. The main financial advantages of subsurface water storage compared to surface water storage consist of (1) increased security of the water supply, (2) little land occupation, (3) reductions in costs of water storage and evaporation losses, and (4) during water purification, storing the treated water in the subsurface can decrease the peak factor allowing the facility to have a smaller capacity (Pyne 2005).

In the Netherlands, climate variability has increased the need for water storage as a result of (1) stronger fluctuations in river discharge and groundwater tables and (2) enhanced intrusion of the North Sea along the Dutch coast. Another driver for storing water underground is the increasing demand for drinking water in areas where the government has forced the water supply companies to restrict their groundwater pumping. This is usually enforced in order to prevent the drawdown of groundwater tables, to restore wetlands in selected areas and to prevent seawater intrusion. Moreover, recent benchmarks of drinking water prices are pushing the Dutch water supply companies to lower their costs. Aquifer storage applications, if operating without quality deterioration, are thus becoming an interesting option also in the Netherlands.

Testing of deep infiltration systems in the Netherlands started in the 1930s without much success due to severe clogging of the infiltration wells. Later in the 1970s, a number of water supply companies (PWN, Waternet, Dunea) started integrating deep-well recharge systems as production units with their existing open infiltration systems. Decisive success factors were the prevention of clogging by extra purification of the recharged water and the treatment of clogging by frequent and intensive back-pumping of the wells. Nonetheless, the ideal situation of recovering stored pre-treated drinking water without the need for post-treatment is rarely achieved.

It is not our intention to provide an exhaustive inventory of all deep-well recharge schemes in the Netherlands but to showcase a few selected interventions aiming at different purposes and targeting the entire spectrum of issues these technologies can resolve. From a research perspective, this overview summarizes in a comprehensive manner the most relevant technological and knowledge advancements related to artificial recharge of aquifers. From the practical side, this overview can serve as a valuable reference for countries facing similar issues to quickly identify solutions suitable for their needs and to inquire further using the given references. Special attention is given to interventions which aimed at storing water for drinking water production as this was the objective in the majority of deep-well recharge sites in the Netherlands (Figure 1). To this end, we provide an in-depth assessment of the common hydrogeochemical processes which can affect the abstracted water quality and render it unsuitable for direct drinking water supply. The aquifer types, their mineralogy and the associated water quality problems compare well with other settings abroad where deep-well recharge sites operate. To cover the complete spectrum of Dutch expertise in this field, we also showcase a few selected innovative applications which aimed at irrigation water production, the achievement of environmental benefits, and wastewater disposal.
Figure 1

Deep-well recharge sites in the Netherlands. 1: Bergje Texel, 2: Haren, 3: Ouddorp, 4: Herten, 5: Zoelen, 6: Watervlak, 7: Waalsdorp, 8: Leersum, 9: Langerak, 10: Nieuwegein, 11: Sint Jansklooster, 12: Scheveningen, 13: Castricum, 14: Leiduin, 15: Breeheide, 16: Someren, 17: Ossendrecht, 18: Noardburgum, 19: Zevenbergen, 20: Ridderkerk, 21: Nootdorp, 22: Ovezande, 23: Westland, 24: Dam 9 Geul 1, 25: Toevoersloot.

Figure 1

Deep-well recharge sites in the Netherlands. 1: Bergje Texel, 2: Haren, 3: Ouddorp, 4: Herten, 5: Zoelen, 6: Watervlak, 7: Waalsdorp, 8: Leersum, 9: Langerak, 10: Nieuwegein, 11: Sint Jansklooster, 12: Scheveningen, 13: Castricum, 14: Leiduin, 15: Breeheide, 16: Someren, 17: Ossendrecht, 18: Noardburgum, 19: Zevenbergen, 20: Ridderkerk, 21: Nootdorp, 22: Ovezande, 23: Westland, 24: Dam 9 Geul 1, 25: Toevoersloot.

DRINKING WATER PRODUCTION

Since the early 1970s, numerous deep-well injection and aquifer storage experiments have been carried out in the Netherlands to acquire insight on (1) the feasibility of recharging confined aquifers with surface water through wells and (2) the water quality changes upon interaction of the injected water with the native groundwater and aquifer sediments (Stuyfzand 1998). The intentional use of natural attenuation processes to improve water quality (Maliva & Missimer 2010) has been deliberately studied to improve understanding of them and fully harness their potential in combination with the infiltration or storage of water. The environmental profits of these natural purification methods are many, the most important being a reduced application of undesired chemicals like coagulants, active carbon, ozone, and chlorine.

The target aquifers are usually composed of Pleistocene or Miocene sands with a grain size of 300–500 μm and a permeability of 30–50 m/d. In most cases the native groundwater is fresh, anoxic, calcite saturated and with relatively high concentrations of Fe(II), Mn(II) and ammonium. Most formations contain low amounts of very reactive iron sulfides (pyrite). Calcium carbonate is typically present in marine deposits and Pleistocene sediments while small amounts of iron carbonates that contain manganese and/or magnesium (e.g. siderite, ankerite, rhodochrosite) are often present.

The quality of the abstracted water is a critical factor to consider as it directly affects the feasibility of applications aiming at drinking water production. The evolution of abstracted water quality is very much scheme-dependent (Figure 2). In aquifer storage recovery (ASR) schemes, the first abstracted volumes are usually of good quality but gradually deteriorate as the interface between injected water and native groundwater approaches the ASR well. Aquifer storage transfer recovery (ASTR) schemes offer a longer aquifer passage and abstracted water quality is consequently expected to improve over time. ASTR schemes suffer, however, from an increased clogging rate of the injection and especially abstraction wells. As opposed to single-purpose groundwater abstraction wells, ASR wells offer a decreased well-clogging risk due to flow reversals which prevent or delay the accumulation of clogging material along the borehole wall. Both scheme types have their advantages and a choice should be made after careful evaluation of aquifer mineralogy and expected hydrogeochemical reactions.
Figure 2

Schematic representation of the differences between the two main deep-well recharge schemes related to drinking water production: 1, 2, and 3 depict injection, storage and recovery, respectively, during ASR cycle 1; 4, 5, and 6 depict the three phases during cycle 2 (modified after Stuyfzand et al. 2012).

Figure 2

Schematic representation of the differences between the two main deep-well recharge schemes related to drinking water production: 1, 2, and 3 depict injection, storage and recovery, respectively, during ASR cycle 1; 4, 5, and 6 depict the three phases during cycle 2 (modified after Stuyfzand et al. 2012).

Hydrogeochemical processes affecting water quality

Upon injection of aerobic water, the water quality is mainly altered by oxygen (and nitrate) consumption by sedimentary electron donors like pyrite, sedimentary organic matter, and exchangeable Fe(II), NH4+, and Mn(II). The induced acidity is partly buffered by HCO3- in the groundwater and, if present, by the dissolution of carbonate minerals. Carbonate dissolution further affects the water quality due to the possible release of Fe(II) and Mn(II) in the groundwater. Heavy metal release may also result from cation exchange (Ca2+ in the injected water displacing Fe(II) and Mn(II) from the exchanger). If oxygen is present then the released Fe(II) is further oxidized (homogeneous oxidation) and precipitates in the form of Fe-hydroxide (ferrihydrite). If however the oxygen front is lagging behind, as it progressively happens further from the injection well, then Fe(II) remains dissolved and may contaminate the water that is abstracted from a well downstream of the injection well (i.e. ASTR schemes). The kinetics of the homogeneous Mn(II) oxidation are very slow (Diem & Stumm 1984) resulting in dissolved Mn(II) contamination being more persistent, even in the presence of oxygen.

The oxidation of pyrite may also lead to some mobilization of trace elements, such as As, Co and Ni. As especially is more mobile due to its reduced state (arsenite), which prevents it from sorbing on the existing and newly formed Fe-hydroxides. With continuous injection, the advancing oxygen (and nitrate) fronts convert the dissolving As into its oxidized state (arsenate), which is much less mobile due to preferential sorption on Fe-hydroxides.

The generalized evolution of quality changes in the water abstracted from a well downstream of the injection well (ASTR scheme) is depicted in Figure 3.
Figure 3

Generalized evolution of the quality changes of oxic injection water in an anoxic, calcareous aquifer, as a function of the number of pore flushes of the soil between injection and abstraction wells. The x-axis also gives the retardation factor for species with a negative change, and the leach factor for species with a positive change (after Stuyfzand 1998).

Figure 3

Generalized evolution of the quality changes of oxic injection water in an anoxic, calcareous aquifer, as a function of the number of pore flushes of the soil between injection and abstraction wells. The x-axis also gives the retardation factor for species with a negative change, and the leach factor for species with a positive change (after Stuyfzand 1998).

The evolution of abstracted water quality allows the identification of a sequence of distinct phases, each of them spanning over a number of pore flushes. The gradual leaching of sedimentary electron donors is testified by the gradually decreasing SO4 production and electron acceptor (NO3- and O2) consumption. The removal of reactive phases is also reflected by the decreasing Ca and HCO3- production, indicating a decline in Ca-carbonate dissolution, a process that is normally triggered as a response to oxidation reactions and the induced acidity. As observed in Figure 3, the complete leaching of calcite requires a much longer flushing time as, upon complete leaching of electron donors, calcite dissolution is only triggered by the negative calcite saturation index of the injected water.

The situation differs in ASR applications where the water is abstracted back from the same dual-purpose well. Iron and manganese which become mobilized during injection (due to the O2 front lagging behind) may show a retarded breakthrough during recovery due to adsorption in the oxidized zone around the ASR well. Adsorption takes place on the newly formed iron hydroxides and on the original cation-exchangers (Figure 4), mainly composed of sedimentary organic material and clay minerals (Appelo et al. 1999). As recovery of the injected water volume continues, Fe(II) and Mn(II) desorb from the outer sorption zone due to decreasing pH conditions and eventually reach the ASR well. Mn(II) especially has a higher tendency to desorb due to its higher pH requirements (pH > 7.5, Buamah et al. 2008) for remaining adsorbed on the exchange sites of the Fe-hydroxide surfaces. In the presence of Mn-containing carbonates, drinking water standards with respect to manganese may be exceeded even sooner due to additional dissolution of these minerals during the early stages of recovery (Antoniou et al. 2012).
Figure 4

ASR scheme with the O2 front lagging behind during the injection phase and dissolved Fe(II) and Mn(II) being adsorbed on ferrihydrite during the early stages of recovery (after Van Der Laan 2009).

Figure 4

ASR scheme with the O2 front lagging behind during the injection phase and dissolved Fe(II) and Mn(II) being adsorbed on ferrihydrite during the early stages of recovery (after Van Der Laan 2009).

Measures to prevent contamination

Field observations on ASR sites and reactive transport modelling have suggested an increasing sorption capacity with subsequent injection–recovery cycles due to a gradual build-up of Fe-hydroxide precipitates. Moreover, a buffer zone can be implemented around the ASR well by halting recovery and restarting injection as soon as drinking water standards are exceeded (Antoniou et al. 2015). This approach shifts the reactive zone further away from the ASR well and can substantially increase the recovery efficiency (i.e. ratio between recovered and injected volume) of the ASR plant. However, if Mn-containing carbonates are present then aquifer pre-treatment or source water pre-treatment may be required to prevent the dissolution of these carbonates during recovery phases and the early exceedance of drinking water guidelines with respect to Fe(II) or Mn(II).

The use of pH-buffering agents such as NaOH and Na2CO3 can increase the alkalinity and pH of the aquifer around the ASR well and prevent the dissolution of carbonate minerals as a response to under-saturated source waters or acidifying water–aquifer interactions. Additionally, increased pH will enhance the sorption of Fe(II) and especially Mn(II) on the Fe-hydroxides formed during injection. The effects of dosing a pH-buffer have been experimentally tested with positive results in real-scale ASR pilots in Virginia and South Carolina which suffered from Fe(II) and Mn(II) concentration exceedances in the recovered water (Ibison et al. 1995; Pyne et al. 2013). Using pH-buffers to increase the aquifer's pH should be avoided if As concentrations are high as adsorbed As (as arsenate) tends to desorb with increasing pH.

Pre-treating the aquifer material was tested by means of column experiments where ASR was realistically simulated in an anoxic setting using real aquifer sediments (Antoniou et al. 2014). It was concluded that treating the aquifer sediments with permanganate can prevent the mobilization of Fe(II) and Mn(II) and greatly improve the recovery efficiency. This is the combined result of (1) the substantial depletion of sedimentary electron donors due to increased oxidation, (2) the extended precipitation of Mn-oxides (by-product of the oxidation reactions) with a high sorption capacity and (3) the increased pH conditions due to proton consumption. Care should be taken so as to prevent the reduction of the Mn-oxides by the inflowing native groundwater during an extended recovery phase or during intentional or circumstantial downtime of the plant, as this may lead to significantly high concentrations of dissolved Mn(II) in the groundwater.

It is expected that ASTR schemes suffering from heavy metal contamination would also benefit from application of pH-buffers although such tests have not been performed to our knowledge. Neutralization of the aquifer material lying between injection and abstraction wells is also beneficial; it should, however, not be performed using permanganate as this would cause severe increases in Mn(II) concentrations in the abstracted water due to the eventual reduction of the newly formed Mn-oxides. A slower neutralization process by injecting water enriched with O2 would be more appropriate.

DEEP-WELL RECHARGE FOR OTHER PURPOSES

Irrigation water production

Subsurface storage of water can also be performed for irrigation water supply. In this case, concentration thresholds of total dissolved solids are less strict than the ones applied for drinking water. This allows brackish or saline aquifers also to be considered for storage and later recovery during high irrigation water demand. A number of recent field experiments in the Netherlands proved that aquifer storage of freshwater in brackish or saline aquifers can be an efficient technique to bridge freshwater shortages in coastal areas (Figure 1). The main challenge in such a setting is dealing with the buoyancy effects which may cause salinization at the bottom of the ASR well during recovery, making part of the fresh water irrecoverable.

A recent study (Zuurbier et al. 2014) was performed in an area dominated by greenhouse horticulture where a high irrigation water demand was until recently covered by storage of rainwater in basins/tanks, use of surface water, and desalination of brackish groundwater (Paalman et al. 2012). The experiment aimed at reducing freshwater losses by applying deep injection and shallow recovery using multiple, individually controlled well screens. This configuration slowed down salinity build-up in the recovery well screens and the recovery efficiency was substantially increased compared to a conventional fully penetrating well.

Achieving environmental goals

Deep-well injection has also been applied in the Netherlands in order to achieve specific environmental goals. In these cases, specific aquifers are usually recharged in order to recover declined water tables to older (higher) levels or to fight salinization due to seawater intrusion.

In 1994 the Amsterdam Water Supply started an experiment to investigate the option of a deep-well recharge system in the Amsterdam dune catchment area, to expand the open recharge system with basins, which has been in operation since 1957 (Olsthoorn & Mosch 2002; Van Duijvenbode & Olsthoorn 2002). Artificial recharge is needed to combat overexploitation of fresh groundwater in the area since the beginning of the 20th century, which has caused groundwater table declines, upconing of brackish/saline water and the disappearance of natural groundwater-dependent vegetation. Pre-treated Rhine river water was injected by four wells screened in the semi-confined aquifer at 30 to 60 metres below mean sea level. The water was recovered by a series of extraction wells 400 metres away. Prior to injection, the already pre-treated (by coagulation and rapid sand filtration) Rhine river water was additionally slowly filtered through fine Aeolian dune sand (Figure 5). This filtration process served to reduce the Membrane Filtration Index (measuring the clogging rate of a 0.45 μm membrane filter – indicative of the physical quality of the water and its ability to be injected) which, especially in summer, rose due to biological activity in the open supply channel.
Figure 5

The deep injection system in the Amsterdam dune catchment area (after Van Duijvenbode & Olsthoorn 2002). 1: pre-treated river water, 2: fine Aeolian dune sand that polishes the water to make it suitable for injection, 3: nylon fabric, 4: gravel pack, 5: drainage system, 6: impermeable plastic sheet to prevent the inflow of the local iron-rich groundwater.

Figure 5

The deep injection system in the Amsterdam dune catchment area (after Van Duijvenbode & Olsthoorn 2002). 1: pre-treated river water, 2: fine Aeolian dune sand that polishes the water to make it suitable for injection, 3: nylon fabric, 4: gravel pack, 5: drainage system, 6: impermeable plastic sheet to prevent the inflow of the local iron-rich groundwater.

The sand filter was cleaned once a year to limit the development of hydraulic resistance due to the continuous infiltration of water with moderate quality. Hydraulic resistance developed also in the injection wells mainly due to the accumulation of biomass and iron flocs in the gravel pack. This resistance disappeared with automatic backwashing, switching on as soon as the water level inside the well exceeded a certain threshold. The removal of clogging which developed on the borehole wall (chemical clogging) required occasional well-surging using compressed air. Currently the extra infiltration capacity in the dune aquifers is not needed and the well infiltration system has stopped.

Disposing wastewater

The production of fresh drinking water from brackish groundwater by reverse osmosis (BWRO) is becoming more attractive, even in temperate climates. The drivers in this case may consist of environmental problems like the pollution and salinization of aquifers, drawdown of water tables in phreatic aquifers, effects of climate change like reduced base flows that render surface waters less fit for drinking water production, and increasing costs to produce drinking water from heavily polluted, fresh groundwater.

In the Netherlands, the injection of the waste saline concentrate into a more saline, confined aquifer is considered to be an ideal disposal solution for this reverse osmosis by-product. An aquifer should meet a number of hydraulic and geochemical requirements in order to qualify as a suitable disposal aquifer (Stuyfzand & Raat 2009). The native groundwater should preferably have a lower quality (in terms of salinity and critical compounds) than the injected concentrate. The BWRO concept is based on the Freshkeeper® approach (Grakist et al. 2002; Kooiman et al. 2004), which stops and reverses the salinization of aquifers and water wells by intercepting the intruding brackish groundwater, thus preventing the underground mixing of fresh and brackish water (Figure 6).
Figure 6

Schematic of (from left to right) a fresh well salinizing by upconing, the fresh-keeper without reverse osmosis (RO), the Freshkeeper with RO, and brackish water with RO (after Stuyfzand & Raat 2009).

Figure 6

Schematic of (from left to right) a fresh well salinizing by upconing, the fresh-keeper without reverse osmosis (RO), the Freshkeeper with RO, and brackish water with RO (after Stuyfzand & Raat 2009).

Two pilots were initiated in 2009 to test the BWRO concept in the Netherlands (Noardburgum and Zevenbergen). The simultaneous abstraction of upper fresh and lower brackish groundwater led to a lowering of the fresh–brackish water interface confirming that instead of lowering production, brackish groundwater should be pumped and used (Raat et al. 2012). At both locations, concentrate injection was technically feasible, as long as the RO recovery levels were not higher than 50% (Zevenbergen) or 70% (Noardburgum). At higher levels, clogging of the injection well due to mineral precipitation could become an issue.

DISCUSSION

From a global perspective, deep-well recharge has become a popular option to provide a reliable water supply. The number of sites has been increasing (Figure 7) with the majority of sites located in the USA, followed by Australia, China, and the Netherlands. As an indication, the number of deep-well recharge wells in the USA alone was 1,200 in 2009, four times higher compared with 1999 (EPA 2016). Despite this recent increase, only a small fraction of projects are located in arid and semi-arid regions even though these regions would greatly benefit from such interventions.
Figure 7

Locations of deep-well recharge sites compiled in the global managed aquifer recharge (MAR) inventory (Stefan 2015) as illustrated in IGRAC's MAR portal (Stefan & Ansems 2016).

Figure 7

Locations of deep-well recharge sites compiled in the global managed aquifer recharge (MAR) inventory (Stefan 2015) as illustrated in IGRAC's MAR portal (Stefan & Ansems 2016).

The majority of sites globally aim at maximizing the natural storage potential of aquifers to eventually enhance domestic water supply (Figure 8). This is not necessarily true anymore in the Netherlands where interventions aiming at alternative goals are becoming progressively more popular. As demonstrated at a number of sites, huge potential exists in the exploitation of brackish and saline aquifers for storage of irrigation water in many coastal areas. Sites operating in a conjunctive manner whereby fresh water is produced from brackish/saline groundwater and the brine is disposed in deeper, low-quality aquifers offer great potential in regions where sustainable solutions are necessary. Prospective developments in deep-well recharge in the Netherlands reach far beyond conventional drinking water production applications. Replicability of the examples could be useful for improving the operation of existing sites abroad as well as for the design and implementation of new projects.
Figure 8

Main objective and final water use of all known deep-well recharge sites compiled in the global MAR inventory (Stefan 2015).

Figure 8

Main objective and final water use of all known deep-well recharge sites compiled in the global MAR inventory (Stefan 2015).

The first step when it comes to replicating these technologies is to assess and map their spatial feasibility for the identification of suitable locations. Predicted spatial suitability was assessed in the Westland-Oostland region, the Netherlands, through comparison with the measured performance of existing sites in the area (Zuurbier et al. 2013). A complete suitability assessment may consist of the integrated analysis of the geomorphological, hydrogeological, socio-economic, and legal conditions pertaining in the region of interest. Products such as suitability maps and information management systems can serve as guidance for interested stakeholders and decision makers in assessing baseline conditions and choosing the solution which best fits to their needs. Information management systems can also help achieve broader replicability and up-scaling of novel solutions as they allow for an interactive and easy-to-follow description of background methodologies.

Although suitability maps serve as a valuable implementation basis, additional real-scale validation should be carried out to confirm the suitability of the proposed locations. As presented in this paper, the success of applications aiming at drinking water production very much depends on the geochemical processes taking place upon aquifer recharge. The presence and reactivity of mineral phases cannot be adequately assessed a priori without piloting or at least laboratory-testing of the aquifer sediments. Fine-scale hydraulics and hydrogeology may also need to be captured to fully assess feasibility at specific locations. It is therefore recommended to always validate maps further via modelling, laboratory-testing and piloting.

CONCLUSIONS

The Dutch experiences showcased in this overview highlight the value of deep-well recharge interventions into both fresh and saline aquifers. Depending on the scope, there are different issues to deal with and different preventive measures to consider. Applications aiming at drinking water supply must pay special attention to hydrogeochemical interactions between recharged water, native groundwater and aquifer sediments. Applications in brackish or saline aquifers should consider feasibility with regards to buoyancy issues and mixing effects. All applications, regardless of purpose and groundwater quality, should consider the possible lateral movement of the injected water caused by the native hydraulic gradient. If the recharged water is to be recovered at a later stage, which is usually the case in aquifer storage applications, this lateral drift may complicate abstraction of the stored water and substantially reduce the recovery efficiency. Spatial suitability maps and information management systems can facilitate the identification of suitable locations and intervention types depending on specific needs.

ACKNOWLEDGEMENTS

The authors would like to thank two anonymous reviewers for their comments and suggestions which improved the quality of this article.

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Bakker
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Zaadnoordijk
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