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Review

Potential Impacts of Induced Bank Filtration on Surface Water Quality: A Conceptual Framework for Future Research

by
Mikael Gillefalk
1,*,
Gudrun Massmann
2,
Gunnar Nützmann
1,3 and
Sabine Hilt
1,*
1
Leibniz-Institute of Freshwater Ecology and Inland Fisheries, Müggelseedamm 301 and 310, 12587 Berlin, Germany
2
Institute of Biology and Environmental Sciences, School of Mathematics and Science, Carl von Ossietzky Universität Oldenburg, Uhlhornsweg 84, 26111 Oldenburg, Germany
3
Geography Department, Faculty of Mathematics and Natural Sciences, Humboldt University Berlin, Unter den Linden 6, 10099 Berlin, Germany
*
Authors to whom correspondence should be addressed.
Water 2018, 10(9), 1240; https://doi.org/10.3390/w10091240
Received: 13 July 2018 / Revised: 4 September 2018 / Accepted: 8 September 2018 / Published: 13 September 2018
(This article belongs to the Section Water Quality and Contamination)
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Abstract

:
Studies on induced bank filtration (IBF), a cost-effective and reliable drinking water production method, usually focus on processes affecting the target drinking water quality. We aim to expand this view by assessing potential impacts of IBF on surface water quality. We suggest that IBF can directly and indirectly affect several physical, chemical and biological processes in both the sediment and open water column, eventually leading to positive or negative changes in source water quality. Direct effects of IBF comprise water level fluctuations, changes in water level and retention time, and in organic content and redox conditions in littoral sediments. Indirect effects are mainly triggered by interrupting groundwater discharge into the surface water body. The latter may result in increased seasonal temperature variations in sediment and water and reduced discharge of solutes transported by groundwater such as nutrients and carbon dioxide. These changes can have cascading effects on various water quality, e.g., by facilitating toxic phytoplankton blooms. We propose investigating these potential effects of IBF in future field and laboratory studies to allow for more detailed insights into these yet unknown effects and their magnitude in order to assure a sustainable application of this valuable technique in the future.

Graphical Abstract

1. Introduction

Bank filtration (BF) is the process by which surface water infiltrates into aquifers. BF occurs when the hydraulic head in the surface water is higher than in the adjacent groundwater. This may naturally be the case, for example in lowland rivers or during high water stages, or caused by groundwater abstraction from wells next to the surface water, a process which is referred to as induced bank filtration (IBF, Figure 1). A stream, lake or river which is subject to BF is also termed a losing stream/lake/river, but water bodies may also be losing in some reaches and gaining in others [1]. It is not uncommon that groundwater abstraction in the vicinity of surface waters leads to unintentional BF, which will be included in the following discussion.
Anthropogenically induced riverbank filtration (RBF) and lake bank filtration (LBF) are alternative ways to assure a potable water supply in sufficient quality and quantity [2]. Other managed aquifer recharge methods are ponded infiltration and soil-aquifer treatment [3]. During IBF suspended solids, bacteria, viruses, parasites or adsorbable or microbially degradable water constituents are partially or fully eliminated (e.g., [4]). IBF has also been recommended as a treatment against odour problems in drinking water gained from surface water [5]. While it is generally widely acknowledged that IBF is a beneficial pre-treatment option for drinking water production, knowledge about its effects on lake and river ecosystems is very limited. Although more than half of the 524 available papers (Web of Science topic search using “*bank filtration”, June 2018) have been classified into Environmental Sciences (Figure 2a), none focuses on the effects of IBF on lake or river ecosystems (one exception being Jacobson et al. [6] that deals with effects on fish habitats). Instead, research concerning IBF has almost exclusively dealt with its purification efficiency as well as infiltration capacity, maintenance considerations and other engineering issues (e.g., [2,4,7,8]). This is surprising, as the source water quality is of high importance for securing high drinking water quality and quantity (Figure 1). In the case of negative effects of IBF on surface water quality, toxic cyanobacteria blooms could occur or be worsened, which would increase the risk of toxin contamination in drinking water even after IBF [9,10] and the need for chlorination [11]. Phytoplankton blooms also increase sedimentation and thus lower hydraulic conductivity and, thereby, the infiltration of surface water into the groundwater [12]. In addition, redox conditions in the sub-surface, which are also affected by surface water quality, can result in increased concentrations of dissolved iron, manganese, hydrogen-sulphide and ammonium in drinking water [4]. We argue that IBF indeed can affect surface water quality and knowledge about this interaction is needed to secure an optimal and sustainable application of this drinking water production technique in the future, avoiding the abandonment of IBF sites as happened in Europe in the last decades [13].
The aims of this study were to (i) assess the extent of IBF usage and the types of surface water bodies potentially affected by IBF by carrying out a literature study on case studies worldwide and to (ii) hypothesize on plausible indirect and direct effects of IBF on surface water bodies in order to (iii) develop a conceptual framework for future research assessing these effects.

2. Use of Induced Bank Filtration (IBF) and Source Surface Water Bodies

2.1. Worldwide Use of IBF and Affected Surface Waters

IBF has been used for more than 100 years and at present is a widely applied method in many European regions [13,14,15] (Table 1). Drinking water derived from infiltrating river and lake water provides a significant share of potable water supplies in various European countries (Table 1). It is also used in North and South America and Asia (Table 1). IBF is not yet utilized in many developing countries, although feasibility studies have been carried out in for example Malawi and Kenya [16].
Most of the published studies on IBF stem from Germany (44%) and the USA (22%, Figure 2b). Germany is the country in Europe with the most IBF sites 46 [13], but studies have been produced in a total of 57 countries (Web of Science topic search using “*bank filtration”, June 2018). According to the Federal Statistical Office of Germany [23], 8.6% of drinking water in Germany originates from IBF, while another 8.8% is defined as “recharged groundwater”, consisting mainly of intentionally recharged surface water. Schmidt et al. [24] state that around 16% of the drinking water in Germany is produced from IBF and other infiltration sites, with more than 300 waterworks using IBF and roughly 50 using artificial groundwater recharge. From a total of 56 studies on IBF mentioning its source water bodies, there were 15 lakes or ponds and 48 different rivers being used for IBF (Table 1). When IBF is conducted along large rivers, such as the Rhine or the Danube, where discharge rates are of a magnitude of 103 m3/s and thus much higher than groundwater influx and abstraction rates (e.g., IBF in Düsseldorf at River Rhine: 0.06% of discharge [73]), the effect of IBF on source water quality is, if not completely negligible, at least very small. In contrast, water quality of lowland rivers, ponds and lakes can potentially be affected by IBF due to their lower discharge (e.g., Lake Müggelsee: up to 50% of discharge, see below).

2.2. Example of IBF Application in Berlin (Germany)

IBF was first applied in Germany’s capital, Berlin, more than 100 years ago. For the past 70 years, bank filtration has produced approximately 60% of the city’s drinking water [4,74]. Water abstraction in Berlin occurs in around 650 wells [75] and is part of a semi-closed water cycle, where effluents from waste water treatment plants reach surface water bodies subject to water extraction via IBF for water provisioning (Figure 3). In total, 9 lakes and many reaches of the lowland rivers Spree, Dahme and Havel are affected by IBF (Figure 3).
In Lake Müggelsee (Figure 4a), mean depth = 4.9 m, surface area = 7.3 km2 [78], groundwater historically discharged into the lake under natural conditions, especially at the northern shore [79] (Figure 4c). However, groundwater withdrawal from well galleries near the shore started in 1905 [78]. Currently, IBF is performed via 170 wells located along the northern, western and southern shore (Figure 4a). Pumping rates are around 40–45 million m3 per year (Figure 4b) distributed among the wells surrounding the lake [80], resulting in a lowering of the groundwater level of by up to 5 m (Figure 4a,c), which is in accordance with groundwater models for the catchment area around Lake Müggelsee [79]. Zippel and Hannappel [74] calculated that 78.4% of the water reaching the abstraction wells was bank filtrate and a substantial part of the lake water was lost via IBF, with the total volume lost each year being almost equal to the volume of the lake (36 million m3 [78]).
During the period 2014–2017 an estimated 1.1 m3/s water was extracted from Lake Müggelsee on average [80], which corresponded to about 28% of the amount of water flowing into Lake Müggelsee via the river Spree within the same period [83,84] (Figure 4b). During periods of low inflow this proportion increased and reached at least 50% on around a fifth of the days, mostly in summer. The amount of groundwater that would reach Lake Müggelsee if no abstraction took place [79] can be estimated by using the pumping rate of the northernmost well galleries (A and B) north of the lake. In the period 2014–2017 the rate was on average 0.4 m3/s [80] and in the same time period the surface water inflow to Lake Müggelsee was 3.9 m3/s [83], which means that the retention time is 10% longer with IBF (107 days compared to 97). Longer retention times also mean a lower flushing rate of nutrients. Comparison of the nutrient retention with and without IBF using the method described by Vollenweider [85] shows that the total phosphorus (P) concentration was 9% higher with IBF (ingoing values: P load = 2000 mg/m2/year [86], discharge = 3.9 m3/s (with IBF) or 4.3 (without IBF) m3/s) compared to if groundwater would reach the lake from the north. These numbers give a first impression of potential impacts by IBF but in order to show detailed effects on lake ecology, field measurements and/or modelling are needed.

3. IBF Effects on Surface Water Quality

Induced bank filtration can potentially affect surface water bodies via two major pathways: (1) directly by induced infiltration of surface water into river and lake sediments; and (2) indirectly by preventing groundwater exfiltration into surface waters. This chapter examines potential IBF effects on physical and chemical parameters affecting biological parameters and processes and eventually surface water quality. In the following paragraphs we first describe the general effects of each respective parameter on water quality and then how IBF potentially changes the parameter.

3.1. Discharge and Retention Time

The flow regime is regarded as key driver of river and floodplain wetland ecosystems [87]. In rivers, low-flow conditions and thus degradation of rivers due to groundwater abstraction and thus unintentional BF has been detected, e.g., in Great Britain [88,89]. Similar problems are expected to be caused by IBF. Lower discharge and water levels in rivers severely change the habitat conditions for macrophytes (e.g., [90]), macroinvertebrates (e.g., [91]) and fish (e.g., [6]). Although the impacts of changes in discharge are manifest across broad taxonomic groups, ecologists still struggle to predict and quantify biotic responses to altered flow regimes [87]. We expect effects on biodiversity, macrophyte abundance, and potentially even the occurrence of harmful cyanobacteria blooms, but the effects depend on the initial conditions (Table 2). Jacobson et al. [6] used a model to show how changing pumping schedules of groundwater extraction well fields could improve the living conditions for fish in a stream by varying water extraction rates and avoiding discharge rates below certain critical thresholds for more than a limited time, thereby ensuring sufficient usable area and discharge.
Higher retention times in lakes have been reported to affect the interaction between phytoplankton and macrophytes. In shallow lakes, this interaction results in the occurrence of alternative stable states with either macrophyte dominance and clear water or phytoplankton-dominated, turbid conditions (see Section 4.2). At higher retention times, this phenomenon is more likely to occur, which has consequences for the management of the waters [92]. Higher retention times also affect nutrient retention, as shown in the Lake Müggelsee case study, where a 10% increase in retention time was estimated to cause a 9% higher total P-concentration (see Section 2.2). In summer, when natural flow is low and water demand is higher, the retention time in lakes naturally increases, hence the effect of IBF is more pronounced then.

3.2. Water Level Fluctuation

Fluctuation of the water level can be a key factor affecting the functioning of lakes (e.g., [109,116]) and rivers (e.g., [104]).
Decreasing water levels may cause former submerged habitats to be exposed to air, resulting in a loss of habitats for littoral plants and animals [117]. Groundwater abstraction was believed to have affected 151 wetland sites of special scientific interest (SSSIs) throughout England and Wales, with 100 additional wetlands perceived to be at future risk [89]. In lakes, water depth and resuspension of sediments due to wind effects are strongly linked [118]. Depending on the lake morphometry, low water levels would enable the wind at a sufficient speed to affect parts of the lake bottom that otherwise would not be affected, especially in shallow lakes. Also, a shallower depth would increase the area of the lake that may be bottom frozen, to which some macrophyte species are sensitive [97]. In Estonia, ground water abstraction starting in 1972 caused lower water levels in three lakes, reaching a decline of up to 3–4 m in the 1980s. This led to the switch from a clear-water to a turbid state in two of the lakes, with subsequent loss of submerged macrophytes. In the 1990s the pumping decreased and the water levels were partially recovered, but the macrophytes did not return [103]. In winter, evergreen macrophytes may become partly fixed in the ice when it forms.
Water-level fluctuations may be an important trigger for the promotion of cyanobacterial blooms ([99] and references therein). In shallow lakes, water level fluctuations may trigger shifts from the clear-water, macrophyte-dominated state to the turbid, cyanobacteria-dominated state (see Section 4.2). In deep reservoirs changing water levels may result in a compressed vertical niche for macrophytes [105] and consequently reduce their inhibiting effects on cyanobacteria [119].
Induced bank filtration has the potential to decrease the water levels of inland waters or increase water level fluctuations in the case of fluctuating pumping regimes (Table 2). These effects are caused directly and indirectly by surface water infiltrating into the sub-surface and by the prevention of groundwater discharge, respectively (Table 2). Various effects of such changes on our target biological parameters have been reported (Table 2). The effects of IBF on surface water quality via effects on water level and its fluctuation strongly depend on a range of additional parameters such as lake morphology, the presence of other regulation measures for water level and background nutrient concentrations. Often, BF-induced water level changes will not be a major factor affecting surface water quality as many lakes and rivers in urban or developed regions are flow- and level regulated.

3.3. Sediment Characteristics

The purification efficiency and infiltration capacity of IBF rely strongly on the characteristics of the sediments where the infiltration of surface water takes place. Therefore, several studies have investigated clogging and redox processes at IBF sites [12,25,26,120]. Most of the purification happens within the first few meters of water transport through the sediment and the following soil layers, although the first few centimetres are already very important for the IBF performance ([14] and references therein) [27,33]. At Lake Tegel (Berlin, Germany), the waterworks pump 45 million m3/year from six wellfields around the lake, and two on the islands in the lake. Pumping from these wells affects hydraulic heads within an area of 50 km2 [121]. Biological clogging with particulate organic matter (POM) reached down to at least 10 cm in the sediments. However, complete clogging did not occur in the sediments, probably due to microbial carbon turnover [25], feeding by detritivorous meiofauna [122] and opening of pores by wave action [26]. The result is a highly active biological zone with algae, bacteria, produced extracellular polymeric substances and meiofauna that still allows for IBF [26]. At low levels of labile organic matter content in the sediments, this process may facilitate macrophyte growth, but with increasing organic matter content the growth rate instead decreases [102]. Low redox potential also negatively impacts certain macrophytes’ propagule emergence [123]. In addition, phosphorus ([124] and references therein) and toxic substances [125] can be released and lead to macrophyte decline and cyanobacteria blooms (Table 2, see also Section 4.1, Section 4.2 and Section 4.3). In surface waters with high sulphate (SO42−) concentrations, e.g., due to lignite mining in the catchment [126], hydrogen sulphide (H2S) might form, which has been shown to be harmful to macrophytes [127].
The infiltration of oxic surface water can also oxidise upper parts of the sediments, especially during periods of high photosynthetic activity of primary producers in the surface water, but studies comparing sediment surface oxygen concentrations with and without IBF are mostly lacking. One exception being Bayarsaikhan et al. [128] who reproduced IBF processes in a laboratory setting and found that degradation of particulate organic carbon (POC) in the form of small pieces of leaf litter consumed almost all oxygen in the sediments.

3.4. Water Temperature

Water temperature is increasingly recognized as an important and highly sensitive driver of water quality [129]. As most biological processes are temperature-dependent, a water temperature regime change can have several consequences for aquatic organisms and their interactions. Temperature directly influences the distribution (e.g., [130,131,132]), predator-prey interactions (e.g., [133]), survival (e.g., [134], growth rates (e.g., [135,136,137,138]), timing of life history events (e.g., [139,140]) and metabolism (e.g., [141,142]) of aquatic organisms in river and lake systems. Immediate effects of higher water temperatures in spring and summer are higher growth rates of primary producers and subsequent changes in their interactions which are crucial for maintaining clear-water conditions (see Section 4.3), as studied in several mesocosm experiments (e.g., [143,144,145,146]). Higher summer water temperatures also promote potentially toxic cyanobacteria blooms, which severely deteriorate water quality (e.g., [147,148,149]). In addition, mineralization of organic matter is fuelled by increasing temperatures and may result in a release of organic-bound P to the sediment pore water [150]. For shallow temperate lakes, Mooij et al. [95] predicted an increased probability of a shift from a clear to a turbid state due to climate change. Their model also predicts higher summer chlorophyll-a concentrations, a stronger dominance of cyanobacteria during summer and reduced zooplankton abundance.
Groundwater usually has a rather stable temperature similar to mean annual air temperatures; for example, groundwater temperatures in Germany vary between 10–12 °C [151]. This fact is often taken advantage of when identifying groundwater discharge zones in surface water by aerial or hand-held thermal infrared (IR) imagery (e.g., [152]). Groundwater infiltrating into surface waters thus has a cooling effect in summer, while it may warm surface waters in winter, as described in Sebok et al. [153]. Another example showed a 3 °C lower water temperature in summer in the part of a lake where groundwater discharged compared to other parts of the same lake [93]. In rivers, groundwater discharge can help keep certain parts free from ice in winter and provide cooling in summer, thereby providing a refuge for fish [135]. Consequently, a change from gaining to losing conditions in a river or lake will result in an increased amplitude of the seasonal temperature variations in surface waters. This may result in a loss of the temperature buffering function of discharging groundwater against surface water freezing in winter and against heating during summer. A lowering of temperatures in winter could lead to increased risk of bottom frozen lakes (see Section 3.2). The effect size on changing temperature buffering depends on the ratio between river discharge and lake volume on the one hand and the abstraction rate on the other. A large river would be less affected than a small river, while for lakes, the size is of less importance since the change would primarily take place in the littoral zone.
A full assessment goes beyond the scope of this study, but effects of IBF-induced temperature increases in summer can be similar to those of climate change-induced warming, especially for benthic organisms in the littoral zone which is highly likely to be warmer with IBF in summer due to the prevention of cold groundwater inflow. Overviews of climate warming effects on lake and river ecosystems are provided by Adrian et al. [94] and Whitehead et al. [154], respectively.

3.5. Nutrient Availability

Although groundwater often has lower nutrient concentrations than surface water, groundwater discharge has been reported as a potential source of additional nutrient (mainly nitrogen and P) loading from anthropogenic sources such as agricultural fertilizers, both in lakes [107,155,156] and rivers [157]. These groundwater-borne nutrients may facilitate the growth of all primary producers and affect their interactions ([107] and references therein), with mostly negative effects on biodiversity and macrophyte abundance (see Section 4.1, Section 4.2 and Section 4.3, Table 2). In oligotrophic systems, submerged macrophytes may either be supported by nutrients from groundwater exfiltration or decline due to competition from periphyton [107,158]. The process of increased nutrient loading to surface waters (eutrophication) often leads to the disappearance of submerged macrophytes [159] and shifts to the turbid state, especially in shallow lakes [160,161] (see Section 4.2), but also lowland rivers [92,162].
Lake Arendsee (Germany) receives more than 50% of its external P loads from groundwater, which significantly adds to its eutrophication [156]. Furthermore, groundwater discharge not only transports nutrients from the catchment into surface waters, but also contributes to a transport of nutrients from sediment pore waters into surface water [158]. Interrupting these loads by installing groundwater wells to induce BF could prevent these nutrients from reaching the surface water and thus contribute to a significant increase in water quality, both through a reduction of the limiting nutrient or by changing the nutrient stoichiometry. In contrast, nutrient-poor groundwater would contribute to a dilution of nutrient-rich water entering lakes or rivers from other sources. Mitigating this dilution effect by inducing BF would not affect the nutrient loading but reduce the actual nutrient concentrations. Furthermore, groundwater is often rich in Fe and Mn which can increase the P-binding capacity. Also, as noted in Section 2.2 and Section 3.1, groundwater exfiltration lowers the retention time in lakes and helps flush the lake of nutrients.

3.6. Pollutants

In general, degradation of contaminants is less efficient in groundwater than in surface water sediment [163]. Antibiotics reaching the groundwater have been shown to change microbial community structure, enhance antibiotic resistance and thereby change ecological functions within the aquifer [164]. Remnants of personal care products and microplastics have been shown to accumulate in sediments [165,166] and to persist there more than in water [167,168]. This accumulation potentially increases by IBF with expected positive effects for pelagic, but negative effects for benthic organisms.
In case the source surface water contains pollutants, more of them could reach the littoral zone and its sediments through IBF, which might facilitate their degradation due to the higher bioactivity, but also change the community structure and abundance of sediment bacteria (e.g., [169]).
In cases where pollutants are transported into surface water by groundwater discharge (e.g., [170]), IBF could interrupt the groundwater flow and thereby stop pollutants from reaching surface water bodies. In cases where concentrations of pollutants are high, groundwater wells would most likely not be installed, but at low concentrations they might be. In such cases, the lake would benefit from pollutant mitigation by IBF.

3.7. Dissolved Inorganic Carbon (DIC) Availability

All aquatic plants can use CO2 as a carbon source [171,172], and since the CO2 concentration needed to half-saturate the photosynthesis of aquatic plants in general is approximately 6–13 times atmospheric levels [173] additional sources are needed. Most lakes of the world (87%) are CO2 supersaturated [174] and CO2 originates from mineralization of organic material in sediments and in the water column, from diffusion from the air but also from surface and groundwater inflow ([175] and references therein). In some cases, groundwater is the sole source of the dissolved inorganic carbon (DIC) influx to lakes ([107] and references therein) and the majority of boreal lakes are CO2-sustained by groundwater [175]. In tropical and temperate lakes, CO2 supersaturation is dependent on groundwater CO2 coming from weathering of minerals [176]. Groundwater in general often contains at least 35 times higher concentrations of CO2 than lakes [177,178], up to 400 times higher in some cases [179,180].
Many macrophyte species have the ability to use HCO3− as a carbon source in addition to free CO2 [181,182]. The use of HCO3− entails higher energetic costs [183]; therefore, macrophytes able to use HCO3− still grow faster in a CO2-rich environment [184]. Other species are fully dependent on CO2 [118,185,186] and as a consequence, CO2 availability is a factor that can control the abundance and species composition of submerged macrophytes. Maberly et al. [187] examined the macrophyte composition along a spring river stretch with a strong gradient of CO2 concentrations dropping from 24 to 5 times the atmospheric concentration. At the headwaters, macrophyte composition was dominated by plants that rely on free CO2 such as the moss Fontinalis antipyretica, whereas with lower CO2 concentrations, the macrophyte composition changed to include more plants that are able to use HCO3−. Productive lakes often depletes free CO2 in summer due to uptake by phytoplankton. In those lakes, macrophytes depending on free CO2 availability may be restricted to areas with a high organic carbon content in the sediment such as shallow protected bays [185,186].
In general, groundwater-borne DIC is supposed to support macrophyte growth, although few empirical studies are available. Frandsen et al. [188] showed that seeping groundwater increased the DIC supply, enhancing the growth of isoetids and to some extent elodeids inhabiting a groundwater-fed softwater lake. Low pH in spring water increased the growth of Egeria densa by affecting the free CO2 concentration in the water [189].
In groundwater-fed lakes where CO2 rich water enters the littoral zone, IBF would interrupt the added CO2 contribution and thereby decrease the possibility for CO2-dependent macrophytes to survive. This process has been suggested to be involved in the complete disappearance of F. antipyretica from Lake Müggelsee (Figure 4) during the last century [190]. An overall lower availability of DIC in the littoral zone due to IBF preventing groundwater exfiltration into the lake may eventually lead to an overall loss of macrophytes. This process may be accelerated by shading through planktonic and periphytic algae, which are able to saturate their need for CO2 at low concentrations of free CO2 due to their small size and effective carbon concentrating mechanisms. Consequently, they are able to grow fast, facing less competition for limiting nutrients with macrophytes [177,191].
Usually, lakes with a high abundance of macrophytes are characterized by a higher biodiversity [98], and a loss of macrophytes would thus potentially lead to a decline in biodiversity (Table 2).

3.8. Dissolved Organic Carbon (DOC)

Groundwater can contain considerable amounts of dissolved organic carbon (DOC) and be responsible for a significant share of the DOC flux into lakes and rivers ([107] and references therein). Natural DOC in shallow groundwater is mainly derived from decomposing organic matter in the soil and often colours the water yellow/brown [192]. Inputs of terrestrial DOC to surface waters have increased in many north temperate and boreal regions over the past decades [112,193]. This browning has several consequences for recipient aquatic ecosystems. The attenuation of light by coloured DOC restricts the growth of benthic primary producers [111,112,113,114]. Williamson et al. [111] also reported fundamental changes in vertical habitat gradients and food web structure in a long-term study on browning in lakes.
Humic substances pose significant challenges during the processing of drinking water supplies ranging from unpleasant taste, odour and colour, to the formation of potentially harmful disinfection by-products when subjected to raw water processing, which often includes treatment with reactive species such as free chlorine, ozone, chloramines, or chlorine dioxide [194]. Since IBF reduces or fully prevents groundwater discharge into inland waters (see e.g., example of Lake Müggelsee), it should thus reduce browning and all changes in water quality affected by it (Table 2). IBF that lowers browning is thus assumed to have a positive effect on biodiversity and macrophyte abundance, while both positive and negative effects on harmful cyanobacteria blooms due to changing light availability and stratification patterns have been reported (Table 2).

4. Summary on IBF Effects on Surface Water Quality

In this chapter we summarize the mechanisms by which IBF can affect biodiversity, macrophyte abundance and cyanobacteria blooms and explain the importance of these surface water quality parameters.

4.1. Biodiversity

Biodiversity of inland waters has been recognized as an invaluable parameter of the ecological quality of inland waters. In European Union (EU) member states, this recognition has resulted in the implementation of the EU Water Framework Directive in 2000. It represents a radical shift towards measuring the status of all surface waters using a range of biological communities rather than the more limited aspects of chemical quality or targeted biological components [195]. In total, 297 assessment methods have been developed by 28 member states with more than half of the methods being based on macroscopic plants (28%) or benthic invertebrates (26%), with the remainder assessing phytoplankton (21%), fish (15%) and phytobenthos (10%) [196]. Jeppesen et al. [197] observed a significant decline in the species richness of zooplankton and submerged macrophytes with increasing total phosphorus (TP) concentrations in the water; while for fish, phytoplankton and floating-leaved macrophytes, species richness was unimodally related to TP, all peaking at 0.1–0.4 mg P/L.
IBF can potentially affect the diversity of all these components of the biological community via its different influences on physical and chemical parameters (Figure 5). In principle, a negative influence on biodiversity seems possible for fish via discharge reductions (see Section 3.1), for macrophytes via water level fluctuations (see Section 3.2) modification of sediment characteristics (see Section 3.3) and reduced CO2 availability (see Section 3.7) and for fish, plants and phytobenthos via the increased temperature amplitude (see Section 3.4) (Table 2). However, IBF could also have a positive effect by reducing the loading by groundwater-born nutrients, DOC and pollutants, thereby increasing biodiversity of certain organism groups.

4.2. Macrophyte Abundance

Submerged macrophytes play a key role for controlling water quality, both in lakes and rivers [92,160,162]. Due to a variety of stabilising mechanisms, a positive feedback between water clarity and macrophytes occurs, especially in shallow lakes and lowland rivers [92,160]. These are often either characterised by macrophyte-dominated conditions with clear water or by phytoplankton dominance and a strong risk of harmful cyanobacteria blooms. Abrupt shifts between these states can be triggered either by changes in nutrient loading or by strong perturbations of their biological structure such as by macrophyte mowing [160,198]. Significantly positive effects of macrophyte stands on water clarity have also been shown for deeper lakes [119,199].
IBF effects on macrophyte abundance may be negative or positive (Figure 5): Macrophyte abundance may be negatively affected by a reduction of CO2 availability (see Section 3.7) and changed redox conditions in the sediments (see Section 3.3) (Table 2). Any negative effects of IBF on macrophytes reduce their inhibiting effects on phytoplankton and thus lower critical threshold levels of shallow lakes for nutrient loading inducing regime shifts to turbid states. They would also reduce their positive effects on mineralization of organic material in the littoral sediments and thus potentially increase the risk of clogging. In contrast, there are also potentially positive effects of IBF on macrophyte abundance, e.g., in water bodies where groundwater discharge is a major source of nutrients, toxic substances inhibiting macrophyte growth, and/or coloured DOC (see Section 3.5, Section 3.6 and Section 3.8) (Table 2).

4.3. Harmful Cyanobacteria Blooms

Cyanobacteria blooms are one of the most severe water quality problems in freshwater ecosystems, especially in lakes and reservoirs [108]. Several species of cyanobacteria produce a wide range of toxic compounds [200]. The incidence and intensity of cyanobacteria blooms and the economic losses associated with these events have increased in recent decades [147,201,202]. The abatement and control of cyanobacteria that produce toxins and create taste and odour problems in drinking water sources is a major challenge for water supplies (e.g., [5,11]). IBF has been shown potential to effectively remove cyanobacteria during underground passage [203]; however, a recent study indicates the potential for the passage of cells even for filamentous cyanobacterial species ([10] and references therein).
Several factors maintaining cyanobacteria blooms in inland waters are potentially facilitated by IBF such as lower flow, higher retention times (see Section 3.1), lower water levels (see Section 3.2), higher summer water temperatures (see Section 3.4), loss of submerged macrophytes (see Section 4.2) and the prevention of groundwater influx containing humic substances inhibiting cyanobacteria blooms (Figure 5). In contrast, the interruption of nutrient-rich groundwater discharge (see Section 3.5) and DOC to surface water by IBF may combat cyanobacteria blooms (Table 2).

5. Conclusions

IBF results in water abstraction from a variety of surface waters worldwide, including ponds, shallow and deep lakes and rivers of different discharge. Being a useful cost-effective and reliable drinking water production method, available studies on IBF only focus on the processes affecting the target drinking water quantity and quality, while its effect on surface waters so far has been ignored.
  • We suggest that IBF directly and indirectly affects physical, chemical and biological processes in surface water that may have both negative and positive effects on their water quality (Figure 5). Potential adverse effects would in turn negatively affect the quality of the water abstracted for drinking water production via IBF (Figure 1 and Figure 5). We predict that IBF-induced changes in water temperature, CO2 availability and water retention times in lakes can lead to macrophyte disappearance, phytoplankton dominance and more suitable conditions for cyanobacteria blooms, among other consequences.
  • Effects of IBF on surface water bodies are assumed to be highest in cases where discharge or lake volumes are small relative to the amount of water abstracted by IBF.
  • Our conceptual impact assessment indicates the need for specific research on IBF effects on source aquatic ecosystems. While field and laboratory experiments may be suitable to test for selected processes, whole ecosystem experiments, monitoring, long-term data sets on aquatic ecosystems before and after the onset of IBF, and modelling are needed to understand the joint impact of IBF.
  • Global change and urbanization are expected to increase the number of surface water bodies being used for IBF. Research on how to minimize potential negative impacts of IBF on their source surface water is thus urgently needed to ensure a sustainable use of this valuable technology.

Author Contributions

M.G. and S.H. conceived the concept, wrote the manuscript and produced figures and tables. G.M. and G.N. participated in the writing and provided feedback on figures and tables. All authors have approved the final article.

Funding

This research was funded by the German Research Foundation grant number GRK 2032/1 (DFG research training group “Urban water interfaces”). The publication of this article was funded by the Open Access Fund of the Leibniz Association.

Acknowledgments

The authors would like to thank the members of the DFG (German Research Foundation) Research Training Group “Urban Water Interfaces”, Eckhard Scheffler, the Berlin Senate Department for the Environment, Transport and Climate Protection, especially Antje Köhler, Alexander Limberg, Benjamin Creutzfeldt and Daniel Brüggemann, for providing data and discussions and the Berlin Water Utilities, especially Dörte Siebenthaler, for her invaluable help during data collection. We are thankful for the linguistic improvements made by Kyle Pipkins and Benjamin Kraemer.

Conflicts of Interest

The authors declare no conflicts of interest.

References

  1. Winter, T.C. Ground Water and Surface Water: A Single Resource; U.S. Geological Survey: Denver, CO, USA, 1998; ISBN 978-0-607-89339-7.
  2. Ray, C.; Melin, G.; Linsky, R.B. (Eds.) Riverbank Filtration: Improving Source-Water Quality; Water Science and Technology Library; Kluwer Acad. Publ: Dordrecht, The Netherlands, 2002; ISBN 978-1-4020-1133-7. [Google Scholar]
  3. Bouwer, H. Artificial recharge of groundwater: Hydrogeology and engineering. Hydrogeol. J. 2002, 10, 121–142. [Google Scholar] [CrossRef]
  4. Hiscock, K.M.; Grischek, T. Attenuation of groundwater pollution by bank filtration. J. Hydrol. 2002, 266, 139–144. [Google Scholar] [CrossRef][Green Version]
  5. Chorus, I.; Klein, G.; Fastner, J.; Rotard, W. Off-flavors in surface waters—How efficient is bank filtration for their abatement in drinking-water. Water Sci. Technol. 1992, 25, 251–258. [Google Scholar] [CrossRef]
  6. Jacobson, R.A.; Warner, G.S.; Parasiewicz, P.; Bagtzoglou, A.C.; Ogden, F.L. An interdisciplinary study of the effects of groundwater extraction on freshwater fishes. Int. J. Ecol. Econ. Stat. 2008, 12, 7–25. [Google Scholar]
  7. Schubert, J. Hydraulic aspects of riverbank filtration—Field studies. J. Hydrol. 2002, 266, 145–161. [Google Scholar] [CrossRef]
  8. Umar, D.A.; Ramli, M.F.; Aris, A.Z.; Sulaiman, W.N.A.; Kura, N.U.; Tukur, A.I. An overview assessment of the effectiveness and global popularity of some methods used in measuring riverbank filtration. J. Hydrol. 2017, 550, 497–515. [Google Scholar] [CrossRef]
  9. Lahti, K.; Rapala, J.; Kivimaki, A.L.; Kukkonen, J.; Niemela, M.; Sivonen, K. Occurrence of microcystins in raw water sources and treated drinking water of Finnish waterworks. Water Sci. Technol. 2001, 43, 225–228. [Google Scholar] [CrossRef] [PubMed]
  10. Pazouki, P.; Prévost, M.; McQuaid, N.; Barbeau, B.; de Boutray, M.-L.; Zamyadi, A.; Dorner, S. Breakthrough of cyanobacteria in bank filtration. Water Res. 2016, 102, 170–179. [Google Scholar] [CrossRef] [PubMed]
  11. Zamyadi, A.; MacLeod, S.L.; Fan, Y.; McQuaid, N.; Dorner, S.; Sauve, S.; Prevost, M. Toxic cyanobacterial breakthrough and accumulation in a drinking water plant: A monitoring and treatment challenge. Water Res. 2012, 46, 1511–1523. [Google Scholar] [CrossRef] [PubMed]
  12. Massmann, G.; Nogeitzig, A.; Taute, T.; Pekdeger, A. Seasonal and spatial distribution of redox zones during lake bank filtration in Berlin, Germany. Environ. Geol. 2008, 54, 53–65. [Google Scholar] [CrossRef]
  13. Sprenger, C.; Hartog, N.; Hernández, M.; Vilanova, E.; Grützmacher, G.; Scheibler, F.; Hannappel, S. Inventory of managed aquifer recharge sites in Europe: Historical development, current situation and perspectives. Hydrogeol. J. 2017, 25, 1909–1922. [Google Scholar] [CrossRef]
  14. Tufenkji, N.; Ryan, J.N.; Elimelech, M. The promise of bank filtration. Environ. Sci. Technol. 2002, 36, 422A–428A. [Google Scholar] [CrossRef] [PubMed]
  15. Ray, C.; Schubert, J.; Linsky, R.B.; Melin, G. Introduction. In Riverbank Filtration: Improving Source-Water Quality; Ray, C., Melin, G., Linsky, R.B., Eds.; Water Science and Technology Library; Kluwer Acad. Publ: Dordrecht, The Netherlands, 2002; pp. 1–15. ISBN 978-1-4020-1133-7. [Google Scholar]
  16. Sharma, S.K.; Amy, G.L. Bank filtration: A sustainable water treatment technology for developing countries. In Water, Sanitation and Hygiene: Sustainable Development and Multisectoral Approaches; Water, Engineering and Development Centre, Loughborough University of Technology, WEDC: Addis Ababa, Ethopia, 2009. [Google Scholar]
  17. Wett, B.; Jarosch, H.; Ingerle, K. Flood induced infiltration affecting a bank filtrate well at the River Enns, Austria. J. Hydrol. 2002, 266, 222–234. [Google Scholar] [CrossRef]
  18. Schön, M. Systematic Comparison of Riverbank Filtration Sites in Austria and India; Leopold Franzens Universität Innsbruck: Innsbruck, Austria, 2014. [Google Scholar]
  19. ICPDR. Groundwater Guidance; International Commission for the Protection of the Danube River: Vienna, Austria, 2016. [Google Scholar]
  20. Miettinen, I.; Martikainen, P.; Vartiainen, T. Humus transformation at the bank filtration water-plant. Water Sci. Technol. 1994, 30, 179–187. [Google Scholar] [CrossRef]
  21. Miettinen, I.T.; Vartiainen, T.; Martikainen, P.J. Bacterial enzyme activities in ground water during bank filtration of lake water. Water Res. 1996, 30, 2495–2501. [Google Scholar] [CrossRef]
  22. Doussan, C.; Ledoux, E.; Detay, M. River-groundwater exchanges, bank filtration, and groundwater quality: Ammonium behavior. J. Environ. Qual. 1998, 27, 1418–1427. [Google Scholar] [CrossRef]
  23. FSO. Public Water Supply; Public Water Supply and Wastewater Treatment; Federal Statistical Office: Wiesbaden, Germany, 2013; p. 18. [Google Scholar]
  24. Schmidt, C.F.; Lange, F.T.; Brauch, H.-J. Assessing the impact of different redox conditions and residence times on the fate of organic micropollutants during riverbank filtration. In Proceedings of the 4th International Conference on Pharmaceuticals and Endocrine Disrupting Chemicals in Water, Minneapolis, MN, USA, 13–15 October 2004; National Ground Water Association: Minneapolis, MN, USA, 2004; pp. 195–205. [Google Scholar]
  25. Hoffmann, A.; Gunkel, G. Carbon input, production and turnover in the interstices of a Lake Tegel bank filtration site, Berlin, Germany. Limnologica 2011, 41, 151–159. [Google Scholar] [CrossRef]
  26. Hoffmann, A.; Gunkel, G. Bank filtration in the sandy littoral zone of Lake Tegel (Berlin): Structure and dynamics of the biological active filter zone and clogging processes. Limnologica 2011, 41, 10–19. [Google Scholar] [CrossRef]
  27. Heberer, T.; Massmann, G.; Fanck, B.; Taute, T.; Duennbier, U. Behaviour and redox sensitivity of antimicrobial residues during bank filtration. Chemosphere 2008, 73, 451–460. [Google Scholar] [CrossRef] [PubMed]
  28. Massmann, G.; Dünnbier, U.; Heberer, T.; Taute, T. Behaviour and redox sensitivity of pharmaceutical residues during bank filtration—Investigation of residues of phenazone-type analgesics. Chemosphere 2008, 71, 1476–1485. [Google Scholar] [CrossRef] [PubMed]
  29. Maeng, S.K.; Ameda, E.; Sharma, S.K.; Grützmacher, G.; Amy, G.L. Organic micropollutant removal from wastewater effluent-impacted drinking water sources during bank filtration and artificial recharge. Water Res. 2010, 44, 4003–4014. [Google Scholar] [CrossRef] [PubMed]
  30. Wiese, B.; Massmann, G.; Jekel, M.; Heberer, T.; Duennbier, U.; Orlikowski, D.; Gruetzmacher, G. Removal kinetics of organic compounds and sum parameters under field conditions for managed aquifer recharge. Water Res. 2011, 45, 4939–4950. [Google Scholar] [CrossRef] [PubMed][Green Version]
  31. Henzler, A.F.; Greskowiak, J.; Massmann, G. Modeling the fate of organic micropollutants during river bank filtration (Berlin, Germany). J. Contam. Hydrol. 2014, 156, 78–92. [Google Scholar] [CrossRef] [PubMed]
  32. Henzler, A.F.; Greskowiak, J.; Massmann, G. Seasonality of temperatures and redox zonations during bank filtration—A modeling approach. J. Hydrol. 2016, 535, 282–292. [Google Scholar] [CrossRef]
  33. Burke, V.; Greskowiak, J.; Asmuss, T.; Bremermann, R.; Taute, T.; Massmann, G. Temperature dependent redox zonation and attenuation of wastewater-derived organic micropollutants in the hyporheic zone. Sci. Total Environ. 2014, 482, 53–61. [Google Scholar] [CrossRef] [PubMed]
  34. Grützmacher, G.; Böttcher, G.; Chorus, I.; Knappe, A.; Pekdeger, A. Cyanobacterial toxins in bank filtered water from Lake Wannsee, Berlin. In Management of Aquifer Recharge for Sustainability; Dillon, P., Ed.; Swets and Zeitlinger: Lisse, The Netherlands, 2002; pp. 175–179. [Google Scholar]
  35. Heberer, T.; Mechlinski, A.; Fanck, B.; Knappe, A.; Massmann, G.; Pekdeger, A.; Fritz, B. Field Studies on the Fate and Transport of Pharmaceutical Residues in Bank Filtration. Groundw. Monit. Remediat. 2004, 24, 70–77. [Google Scholar] [CrossRef]
  36. Massmann, G.; Sueltenfuss, J.; Duennbier, U.; Knappe, A.; Taute, T.; Pekdeger, A. Investigation of groundwater a residence times during bank filtration in Berlin: Multi-tracer approach. Hydrol. Process. 2008, 22, 788–801. [Google Scholar] [CrossRef]
  37. Kohfahl, C.; Massmann, G.; Pekdeger, A. Sources of oxygen flux in groundwater during induced bank filtration at a site in Berlin, Germany. Hydrogeol. J. 2009, 17, 571–578. [Google Scholar] [CrossRef]
  38. Chorus, I. (Ed.) Cyanotoxins: Occurrence, Causes, Consequences; Springer: Berlin, Germany; New York, NY, USA, 2012; ISBN 978-3-642-59514-1. [Google Scholar]
  39. Achten, C.; Kolb, A.; Puttmann, W. Occurrence of methyl tert-butyl ether (MTBE) in riverbank filtered water and drinking water produced by riverbank filtration 2. Environ. Sci. Technol. 2002, 36, 3662–3670. [Google Scholar] [CrossRef] [PubMed]
  40. Ray, C.; Soong, T.W.; Lian, Y.Q.; Roadcap, G.S. Effect of flood-induced chemical load on filtrate quality at bank filtration sites. J. Hydrol. 2002, 266, 235–258. [Google Scholar] [CrossRef]
  41. Homonnay, Z. Use of Bank Filtration in Hungary. In Riverbank Filtration: Understanding Contaminant Biogeochemistry and Pathogen Removal; Ray, C., Ed.; Springer: Dordrecht, The Netherlands, 2002; pp. 221–228. ISBN 978-1-4020-0955-6. [Google Scholar]
  42. Rossetto, R.; Barbagli, A.; Borsi, I.; Mazzanti, G.; Vienken, T.; Bonari, E. Site investigation and design of the monitoring system at the Sant’Alessio Induced RiverBank Filtration plant (Lucca, Italy). Rendiconti online della Società Geologica Italiana 2015, 35, 248–251. [Google Scholar] [CrossRef]
  43. Eynard, F.; Mez, K.; Walther, J.-L. Risk of cyanobacterial toxins in Riga waters (Latvia). Water Res. 2000, 34, 2979–2988. [Google Scholar] [CrossRef]
  44. Medema, G.J.; Stuyfzand, P.J. Removal of micro-organisms upon basin recharge, deep well injection and river bank filtration in the Netherlands. In Management of Aquifer Recharge for Sustainability; Dillon, P., Ed.; Swets and Zeitlinger: Lisse, The Netherlands, 2002; pp. 125–131. ISBN 90 5809 527 4. [Google Scholar]
  45. Hamann, E.; Stuyfzand, P.J.; Greskowiak, J.; Timmer, H.; Massmann, G. The fate of organic micropollutants during long-term/long-distance river bank filtration. Sci. Total Environ. 2016, 545, 629–640. [Google Scholar] [CrossRef] [PubMed]
  46. Mollema, P.N.; Stuyfzand, P.J.; Juhasz-Holterman, M.H.A.; Van Diepenbeek, P.M.J.A.; Antonellini, M. Metal accumulation in an artificially recharged gravel pit lake used for drinking water supply. J. Geochem. Explor. 2015, 150, 35–51. [Google Scholar] [CrossRef]
  47. Mollema, P.N.; Antonellini, M.; Hubeek, A.; Van Diepenbeek, P.M.J.A. The effect of artificial recharge on hydrochemistry: A comparison of two fluvial gravel pit lakes with different post-excavation uses in the netherlands. Water 2016, 8, 409. [Google Scholar] [CrossRef]
  48. Kvitsand, H.M.L.; Myrmel, M.; Fiksdal, L.; Osterhus, S.W. Evaluation of bank filtration as a pretreatment method for the provision of hygienically safe drinking water in Norway: Results from monitoring at two full-scale sites. Hydrogeol. J. 2017, 25, 1257–1269. [Google Scholar] [CrossRef]
  49. Przybylek, J.; Dragon, K.; Kaczmarek, P.M.J. Hydrogeological investigations of river bed clogging at a river bank filtration site along the River Warta, Poland. Geologos 2017, 23, 201–214. [Google Scholar] [CrossRef][Green Version]
  50. Rojanschi, V.; Mlenajek, L.; Stanciulesc, M. Riverbank filtration in water supply in Romania—Old solutions, new problems. In Riverbank Filtration: Understanding Contaminant Biogeochemistry and Pathogen Removal; Ray, C., Ed.; Springer Science & Business Media: Berlin, Germany, 2002; pp. 235–245. ISBN 978-1-4020-0954-9. [Google Scholar]
  51. Vogt, T.; Hoehn, E.; Schneider, P.; Cirpka, O.A. Untersuchung der Flusswasserinfiltration in voralpinen Schottern mittels Zeitreihenanalyse. Grundwasser 2009, 14, 179–194. [Google Scholar] [CrossRef][Green Version]
  52. Castella, E.; Bickerton, M.; Armitage, P.D.; Petts, G.E. The effects of water abstractions on invertebrate communities in U.K. streams. Hydrobiologia 1995, 308, 167–182. [Google Scholar] [CrossRef]
  53. Ahmed, A.K.A.; Marhaba, T.F. Review on river bank filtration as an in situ water treatment process. Clean Technol. Environ. Policy 2017, 19, 349–359. [Google Scholar] [CrossRef]
  54. Weiss, W.J.; Bouwer, E.J.; Aboytes, R.; LeChevallier, M.W.; O’Melia, C.R.; Le, B.T.; Schwab, K.J. Riverbank filtration for control of microorganisms: Results from field monitoring. Water Res. 2005, 39, 1990–2001. [Google Scholar] [CrossRef] [PubMed]
  55. Sahoo, G.B.; Ray, C.; Wang, J.Z.; Hubbs, S.A.; Song, R.; Jasperse, J.; Seymour, D. Use of artificial neural networks to evaluate the effectiveness of riverbank filtration. Water Res. 2005, 39, 2505–2516. [Google Scholar] [CrossRef] [PubMed]
  56. Ray, C. (Ed.) Riverbank Filtration: Understanding Contaminant Biogeochemistry and Pathogen Removal; Springer Science & Business Media: Berlin, Germany, 2002; ISBN 978-1-4020-0954-9. [Google Scholar]
  57. Harvey, R.W.; Metge, D.W.; LeBlanc, D.R.; Underwood, J.; Aiken, G.R.; Butler, K.; McCobb, T.D.; Jasperse, J. Importance of the Colmation Layer in the Transport and Removal of Cyanobacteria, Viruses, and Dissolved Organic Carbon during Natural Lake-Bank Filtration. J. Environ. Qual. 2015, 44, 1413–1423. [Google Scholar] [CrossRef] [PubMed]
  58. Freitas, D.A.; Cabral, J.J.S.P.; Rocha, F.J.S.; Paiva, A.L.R.; Sens, M.L.; Veras, T.B. Cryptosporidium spp. and Giardia spp. removal by bank filtration at Beberibe River, Brazil. River Res. Appl. 2017, 33, 1079–1087. [Google Scholar] [CrossRef]
  59. Romero, L.G.; Mondardo, R.I.; Sens, M.L.; Grischek, T. Removal of cyanobacteria and cyanotoxins during lake bank filtration at Lagoa do Peri, Brazil. Clean Technol. Environ. Policy 2014, 16, 1133–1143. [Google Scholar] [CrossRef]
  60. Romero-Esquivel, L.G.; Pizzolatti, B.S.; Sens, M.L. Potential application of bank filtration in Santa Catarina, Brazil. Interciencia 2016, 41, 740–747. [Google Scholar]
  61. Hu, B.; Teng, Y.; Zhai, Y.; Zuo, R.; Li, J.; Chen, H. Riverbank filtration in China: A review and perspective. J. Hydrol. 2016, 541, 914–927. [Google Scholar] [CrossRef]
  62. Sandhu, C.; Grischek, T.; Kumar, P.; Ray, C. Potential for riverbank filtration in India. Clean Technol. Environ. Policy 2011, 13, 295–316. [Google Scholar] [CrossRef]
  63. Lorenzen, G.; Sprenger, C.; Taute, T.; Pekdeger, A.; Mittal, A.; Massmann, G. Assessment of the potential for bank filtration in a water-stressed megacity (Delhi, India). Environ. Earth Sci. 2010, 61, 1419–1434. [Google Scholar] [CrossRef]
  64. Sharma, B.; Uniyal, D.P.; Dobhal, R.; Kimothi, P.C.; Grischek, T. A sustainable solution for safe drinking water through bank filtration technology in Uttarakhand, India. Curr. Sci. 2014, 107, 1118–1124. [Google Scholar]
  65. Thakur, A.K.; Ojha, C.S.P.; Singh, V.P.; Gurjar, B.R.; Sandhu, C. Removal of pathogens by river bank filtration at Haridwar, India. Hydrol. Process. 2013, 27, 1535–1542. [Google Scholar] [CrossRef]
  66. Dash, R.R.; Mehrotra, I.; Kumar, P.; Grischek, T. Lake bank filtration at Nainital, India: Water-quality evaluation. Hydrogeol. J. 2008, 16, 1089–1099. [Google Scholar] [CrossRef]
  67. Othman, S.Z.; Adlan, M.N.; Selamat, M.R. A study on the potential of riverbank filtration for the removal of color, iron, turbidy and E.coli in Sungai Perak, Kota Lama Kiri, Kuala Kangsar, Perak, Malaysia. Jurnal Teknologi 2015, 74, 83–91. [Google Scholar] [CrossRef]
  68. Bork, J.; Berkhoff, S.E.; Bork, S.; Hahn, H.J. Using subsurface metazoan fauna to indicate groundwater-surface water interactions in the Nakdong River floodplain, South Korea. Hydrogeol. J. 2009, 17, 61–75. [Google Scholar] [CrossRef]
  69. Pholkern, K.; Srisuk, K.; Grischek, T.; Soares, M.; Schäfer, S.; Archwichai, L.; Saraphirom, P.; Pavelic, P.; Wirojanagud, W. Riverbed clogging experiments at potential river bank filtration sites along the Ping River, Chiang Mai, Thailand. Environ. Earth Sci. 2015, 73, 7699–7709. [Google Scholar] [CrossRef]
  70. Ghodeif, K.; Grischek, T.; Bartak, R.; Wahaab, R.; Herlitzius, J. Potential of river bank filtration (RBF) in Egypt. Environ. Earth Sci. 2016, 75, 671. [Google Scholar] [CrossRef]
  71. Shamrukh, M.; Abdel-Wahab, A. Riverbank filtration for sustainable water supply: Application to a large-scale facility on the Nile River. Clean Technol. Environ. Policy 2008, 10, 351–358. [Google Scholar] [CrossRef]
  72. Hamdan, A.M.; Sensoy, M.M.; Mansour, M.S. Evaluating the effectiveness of bank infiltration process in new Aswan City, Egypt. Arab. J. Geosci. 2013, 6, 4155–4165. [Google Scholar] [CrossRef]
  73. Stadtwerke Düsseldorf. Clearly Drinking Water—Element of Life; Stadtwerke Düsseldorf: Düsseldorf, Germany, 2011. [Google Scholar]
  74. Zippel, M.; Hannappel, S. Evaluation of the groundwater yield of Berlin water works using regional numerical groundwater flow models. Grundwasser 2008, 13, 195–207. [Google Scholar] [CrossRef]
  75. Berlin Water Utilities. Groundwater Extraction. Available online: http://www.bwb.de/content/language1/html/961.php (accessed on 1 June 2018).
  76. Berlin Water Utilities. Wasserschutzgebiete auf Einen Blick; Berlin Water Utilities: Berlin, Germany, 2018. [Google Scholar]
  77. Jekel, M.; Ruhl, A.S.; Meinel, F.; Zietzschmann, F.; Lima, S.P.; Baur, N.; Wenzel, M.; Gnirß, R.; Sperlich, A.; Dünnbier, U.; et al. Anthropogenic organic micro-pollutants and pathogens in the urban water cycle: Assessment, barriers and risk communication (ASKURIS). Environ. Sci. Eur. 2013, 25, 20. [Google Scholar] [CrossRef][Green Version]
  78. Driescher, E.; Behrendt, H.; Schellenberger, G.; Stellmacher, R. Lake Müggelsee and its environment—Natural conditions and anthropogenic impacts. Int. Revue Gesamten Hydrobiol. Hydrogr. 1993, 78, 327–343. [Google Scholar] [CrossRef]
  79. Zippel, M. Model-Based Assessment of the Underground Water Resources of Berlin—The Groundwater Situation in the Catchment Area of Lake Müggelsee; Freie Universität: Berlin, Germany, 2006. [Google Scholar]
  80. Berlin Water Utilities. Pumping Rates from Well Galleries around Lake Müggelsee; Berlin Water Utilities: Berlin, Germany, 2018. [Google Scholar]
  81. Schröter, W. Groundwater Levels of the Main Aquifer and Panke Valley Aquifer. 2015. Available online: http://www.stadtentwicklung.berlin.de/umwelt/umweltatlas/kq212.htm (accessed on 15 May 2018).
  82. Frey, W.; Skripalle, J.; Krüger, P. Ecological Consequences of Large-Scale Groundwater Drawdown; Senate of Berlin, Department for the Environment, Transport and Climate Protection: Berlin, Germany, 1992. [Google Scholar]
  83. Senate of Berlin. Measured and Modelled Water Discharge into Lake Müggelsee; Berlin Senate Department for the Environment, Transport and Climate Protection: Berlin, Germany, 2018. [Google Scholar]
  84. Schumacher, F.; Storz, P. State of Model 2016 and Transient Calculation of the Hydrological Years 2004 to 2014; Senate of Berlin, Department for the Environment, Transport and Climate Protection: Berlin, Germany, 2016; 12p. [Google Scholar]
  85. Vollenweider, R.A. Advances in defining critical loading levels for phosphorus in lake eutrophication. Mem. Ist. Ital. Idrobiol. 1976, 33, 53–83. [Google Scholar]
  86. Shatwell, T.; Köhler, J. Decreased nitrogen loading controls summer cyanobacterial blooms without promoting nitrogen-fixing taxa: Long-term response of a shallow lake. Limnol. Oceanogr. 2018. [Google Scholar] [CrossRef]
  87. Bunn, S.E.; Arthington, A.H. Basic principles and ecological consequences of altered flow regimes for aquatic biodiversity. Environ. Manag. 2002, 30, 492–507. [Google Scholar] [CrossRef]
  88. Bickerton, M.; Petts, G.; Armitage, P.; Castella, E. Assessing the ecological effects of groundwater abstraction on chalk streams: Three examples from Eastern England. Regul. Rivers Res. Mgmt. 1993, 8, 121–134. [Google Scholar] [CrossRef]
  89. Acreman, M.C.; Adams, B.; Birchall, P.; Connorton, B. Does groundwater abstraction cause degradation of rivers and wetlands? Water Environ. J. 2000, 14, 200–206. [Google Scholar] [CrossRef]
  90. Hilt, S.; Schoenfelder, I.; Rudnickaa, A.; Carls, R.; Nikolaevich, N.; Sukhodolov, A.; Engelhardt, C. Reconstruction of pristine morphology, flow, nutrient conditions and submerged vegetation of lowland river spree (Germany) from palaeomeanders. River Res. Appl. 2008, 24, 310–329. [Google Scholar] [CrossRef]
  91. Armitage, P.; Petts, G. Biotic score and prediction to assess the effects of water abstractions. Aquat. Conserv. Mar. Freshw. Ecosyst. 1992, 2, 1–17. [Google Scholar] [CrossRef]
  92. Hilt, S.; Koehler, J.; Kozerski, H.-P.; van Nes, E.H.; Scheffer, M. Abrupt regime shifts in space and time along rivers and connected lake systems. Oikos 2011, 120, 766–775. [Google Scholar] [CrossRef]
  93. Liu, C.; Liu, J.; Wang, X.-S.; Zheng, C. Analysis of groundwater–lake interaction by distributed temperature sensing in Badain Jaran Desert, Northwest China. Hydrol. Process. 2016, 30, 1330–1341. [Google Scholar] [CrossRef]
  94. Adrian, R.; Hessen, D.O.; Blenckner, T.; Hillebrand, H.; Hilt, S.; Jeppesen, E.; Livingstone, D.M.; Trolle, D. Environmental impacts—Lake ecosystems. In North Sea Region Climate Change Assessment; Quante, M., Colijn, F., Eds.; Springer International Publ.: Cham, Switzerland, 2016; pp. 315–340. ISBN 978-3-319-39743-6. [Google Scholar]
  95. Mooij, W.M.; Janse, J.H.; Domis, L.N.D.S.; Huelsmann, S.; Ibelings, B.W. Predicting the effect of climate change on temperate shallow lakes with the ecosystem model PCLake. Hydrobiologia 2007, 584, 443–454. [Google Scholar] [CrossRef][Green Version]
  96. Cieśliński, R.; Piekarz, J.; Zieliński, M. Groundwater supply of lakes: The case of Lake Raduńskie Górne (northern Poland, Kashubian Lake District). Hydrol. Sci. J. 2016, 61, 2427–2434. [Google Scholar] [CrossRef]
  97. Hellsten, S. Environmental Factors and Aquatic Macrophytes in the Littoral Zone of Regulated Lakes; University of Oulo: Oulo, Finland, 2000. [Google Scholar]
  98. Hilt, S.; Brothers, S.; Jeppesen, E.; Veraart, A.J.; Kosten, S. Translating Regime Shifts in Shallow Lakes into Changes in Ecosystem Functions and Services. BioScience 2017, 67, 928–936. [Google Scholar] [CrossRef]
  99. Bakker, E.S.; Hilt, S. Impact of water-level fluctuations on cyanobacterial blooms: Options for management. Aquat. Ecol. 2016, 50, 485–498. [Google Scholar] [CrossRef]
  100. Rorslett, B.; Johansen, S.W. Remedial measures connected with aquatic macrophytes in Norwegian regulated rivers and reservoirs. Regul. Rivers Res. Manag. 1996, 12, 509–522. [Google Scholar] [CrossRef]
  101. Mitrovic, S.M.; Hardwick, L.; Dorani, F. Use of flow management to mitigate cyanobacterial blooms in the Lower Darling River, Australia. J. Plankton Res. 2011, 33, 229–241. [Google Scholar] [CrossRef]
  102. Barko, J.W.; Gunnison, D.; Carpenter, S.R. Sediment interactions with submersed macrophyte growth and community dynamics. Aquat. Bot. 1991, 41, 41–65. [Google Scholar] [CrossRef]
  103. Vandel, E.; Vaasma, T.; Terasmaa, J.; Koff, T.; Vainu, M. Effect of human induced drastic water-level changes to ecologically sensitive small lakes. In Proceedings of 2nd International Conference—Water Resources and Wetlands; Romanian Limnogeographical Association: Târgoviște, Romania, 2014; pp. 204–211. [Google Scholar]
  104. Leyer, I. Predicting plant species’ responses to river regulation: The role of water level fluctuations. J. Appl. Ecol. 2005, 42, 239–250. [Google Scholar] [CrossRef]
  105. Rorslett, B. Environmental-factors and aquatic macrophyte response in regulated lakes—A statistical approach. Aquat. Bot. 1984, 19, 199–220. [Google Scholar] [CrossRef]
  106. Stets, E.G.; Striegl, R.G.; Aiken, G.R.; Rosenberry, D.O.; Winter, T.C. Hydrologic support of carbon dioxide flux revealed by whole-lake carbon budgets. J. Geophys. Res. 2009, 114, G01008. [Google Scholar] [CrossRef]
  107. Périllon, C.; Hilt, S. Groundwater influence differentially affects periphyton and macrophyte production in lakes. Hydrobiologia 2015, 1–13. [Google Scholar] [CrossRef]
  108. O’Neil, J.M.; Davis, T.W.; Burford, M.A.; Gobler, C.J. The rise of harmful cyanobacteria blooms: The potential roles of eutrophication and climate change. Harmful Algae 2012, 14, 313–334. [Google Scholar] [CrossRef]
  109. Jeppesen, E.; Meerhoff, M.; Davidson, T.A.; Trolle, D.; Sondergaard, M.; Lauridsen, T.L.; Beklioglu, M.; Brucet, S.; Volta, P.; Gonzalez-Bergonzoni, I.; et al. Climate change impacts on lakes: An integrated ecological perspective based on a multi-faceted approach, with special focus on shallow lakes. J. Limnol. 2014, 73, 88–111. [Google Scholar] [CrossRef][Green Version]
  110. Phillips, G.L.; Eminson, D.; Moss, B. A mechanism to account for macrophyte decline in progressively eutrophicated freshwaters. Aquat. Bot. 1978, 4, 103–126. [Google Scholar] [CrossRef]
  111. Williamson, C.E.; Overholt, E.P.; Pilla, R.M.; Leach, T.H.; Brentrup, J.A.; Knoll, L.B.; Mette, E.M.; Moeller, R.E. Ecological consequences of long-term browning in lakes. Sci. Rep. 2016, 5. [Google Scholar] [CrossRef] [PubMed]
  112. Solomon, C.T.; Jones, S.E.; Weidel, B.C.; Buffam, I.; Fork, M.L.; Karlsson, J.; Larsen, S.; Lennon, J.T.; Read, J.S.; Sadro, S.; et al. Ecosystem Consequences of Changing Inputs of Terrestrial Dissolved Organic Matter to Lakes: Current Knowledge and Future Challenges. Ecosystems 2015, 18, 376–389. [Google Scholar] [CrossRef]
  113. Vasconcelos, F.R.; Diehl, S.; Rodriguez, P.; Hedstrom, P.; Karlsson, J.; Bystrom, P. Asymmetrical competition between aquatic primary producers in a warmer and browner world. Ecology 2016, 97, 2580–2592. [Google Scholar] [CrossRef] [PubMed]
  114. Brothers, S.; Koehler, J.; Attermeyer, K.; Grossart, H.P.; Mehner, T.; Meyer, N.; Scharnweber, K.; Hilt, S. A feedback loop links brownification and anoxia in a temperate, shallow lake. Limnol. Oceanogr. 2014, 59, 1388–1398. [Google Scholar] [CrossRef][Green Version]
  115. Relyea, R.A. The impact of insecticides and herbicides on the biodiversity and productivity of aquatic communities. Ecol. Appl. 2005, 15, 618–627. [Google Scholar] [CrossRef]
  116. Coops, H.; Beklioglu, M.; Crisman, T.L. The role of water-level fluctuations in shallow lake ecosystems—Workshop conclusions. Hydrobiologia 2003, 506–509, 23–27. [Google Scholar] [CrossRef]
  117. Leira, M.; Cantonati, M. Effects of water-level fluctuations on lakes: An annotated bibliography. Hydrobiologia 2008, 613, 171–184. [Google Scholar] [CrossRef]
  118. Scheffer, M. Ecology of Shallow Lakes; Springer: Dordrecht, The Netherlands, 2004; ISBN 978-1-4020-2306-4. [Google Scholar]
  119. Sachse, R.; Petzoldt, T.; Blumstock, M.; Moreira, S.; Paetzig, M.; Ruecker, J.; Janse, J.H.; Mooij, W.M.; Hilt, S. Extending one-dimensional models for deep lakes to simulate the impact of submerged macrophytes on water quality. Environ. Model. Softw. 2014, 61, 410–423. [Google Scholar] [CrossRef]
  120. Wiese, B.; Nützmann, G. Transient leakance and infiltration characteristics during lake bank filtration. Groundwater 2009, 47, 57–68. [Google Scholar] [CrossRef] [PubMed]
  121. WASY. Hydrogeological Report on the Impact of Fundamental Changes in the Operation of Groundwater Enrichment at Tegel Waterworks; WASY Institute for Water Resources Planning and Systems Research Ltd.: Berlin, Germany, 2004. [Google Scholar]
  122. Gunkel, G.; Beulker, C.; Hoffmann, A.; Kosmol, J. Fine particulate organic matter (FPOM) transport and processing in littoral interstices—Use of fluorescent markers. Limnologica 2009, 39, 185–199. [Google Scholar] [CrossRef]
  123. Van Zuidam, B.G.; Cazemier, M.M.; van Geest, G.J.; Roijackers, R.M.M.; Peeters, E.T.H.M. Relationship between redox potential and the emergence of three submerged macrophytes. Aquat. Bot. 2014, 113, 56–62. [Google Scholar] [CrossRef]
  124. Ding, S.; Wang, Y.; Wang, D.; Li, Y.Y.; Gong, M.; Zhang, C. In situ, high-resolution evidence for iron-coupled mobilization of phosphorus in sediments. Sci. Rep. 2016, 6, 24341. [Google Scholar] [CrossRef] [PubMed][Green Version]
  125. Shaheen, S.M.; Rinklebe, J.; Rupp, H.; Meissner, R. Temporal dynamics of pore water concentrations of Cd, Co, Cu, Ni, and Zn and their controlling factors in a contaminated floodplain soil assessed by undisturbed groundwater lysimeters. Environ. Pollut. 2014, 191, 223–231. [Google Scholar] [CrossRef] [PubMed]
  126. Uhlmann, W.; Zimmermann, K. Case Analysis of the Sulphate Load in the River Spree 2014/15; Senate of Berlin, Department for the Environment, Transport and Climate Protection: Dresden, Germany, 2015; 96p. [Google Scholar]
  127. Koch, M.S.; Mendelssohn, I.A.; McKee, K.L. Mechanism for the hydrogen sulfide-induced growth limitation in wetland macrophytes. Limnol. Oceanogr. 1990, 35, 399–408. [Google Scholar] [CrossRef][Green Version]
  128. Bayarsaikhan, U.; Filter, J.; Gernert, U.; Jekel, M.; Ruhl, A.S. Fate of leaf litter deposits and impacts on oxygen availability in bank filtration column studies. Environ. Res. 2018, 164, 495–500. [Google Scholar] [CrossRef] [PubMed]
  129. Hannah, D.M.; Webb, B.W.; Nobilis, F. River and stream temperature: Dynamics, processes, models and implications—Preface. Hydrol. Process. 2008, 22, 899–901. [Google Scholar] [CrossRef]
  130. Boisneau, C.; Moatar, F.; Bodin, M.; Boisneau, P. Does global warming impact on migration patterns and recruitment of Allis shad (Alosa alosa L.) young of the year in the Loire River, France? Hydrobiologia 2008, 602, 179–186. [Google Scholar] [CrossRef]
  131. Yvon-Durocher, G.; Montoya, J.M.; Trimmer, M.; Woodward, G. Warming alters the size spectrum and shifts the distribution of biomass in freshwater ecosystems. Glob. Chang. Biol. 2011, 17, 1681–1694. [Google Scholar] [CrossRef]
  132. Hussner, A.; van Dam, H.; Vermaat, J.E.; Hitt, S. Comparison of native and neophytic aquatic macrophyte developments in a geothermally warmed river and thermally normal channels. Fundam. Appl. Limnol. 2014, 185, 155–165. [Google Scholar] [CrossRef]
  133. Boscarino, B.T.; Rudstam, L.G.; Mata, S.; Gal, G.; Johannsson, O.E.; Mills, E.L. The effects of temperature and predator—Prey interactions on the migration behavior and vertical distribution of Mysis relicta. Limnol. Oceanogr. 2007, 52, 1599–1613. [Google Scholar] [CrossRef]
  134. Wehrly, K.E.; Wang, L.; Mitro, M. Field-based estimates of thermal tolerance limits for trout: Incorporating exposure time and temperature fluctuation. Trans. Am. Fish. Soc. 2007, 136, 365–374. [Google Scholar] [CrossRef]
  135. Power, G.; Brown, R.S.; Imhof, J.G. Groundwater and fish—Insights from northern North America. Hydrol. Process. 1999, 13, 401–422. [Google Scholar] [CrossRef]
  136. Imholt, C.; Gibbins, C.N.; Malcolm, I.A.; Langan, S.; Soulsby, C. Influence of riparian cover on stream temperatures and the growth of the mayfly Baetis rhodani in an upland stream. Aquat. Ecol. 2010, 44, 669–678. [Google Scholar] [CrossRef]
  137. Jensen, A.J. Atlantic salmon (Salmo salar) in the regulated River Alta: Effects of altered water temperature on parr growth. River Res. Appl. 2003, 19, 733–747. [Google Scholar] [CrossRef]
  138. Whitledge, G.W.; Rabeni, C.F.; Annis, G.; Sowa, S.P. Riparian Shading and Groundwater Enhance Growth Potential for Smallmouth Bass in Ozark Streams. Ecol. Appl. 2006, 16, 1461–1473. [Google Scholar] [CrossRef]
  139. Gerten, D.; Adrian, R. Species-specific changes in the phenology and peak abundance of freshwater copepods in response to warm summers. Freshw. Biol. 2002, 47, 2163–2173. [Google Scholar] [CrossRef]
  140. Harper, M.P.; Peckarsky, B.L. Emergence cues of a mayfly in a high-altitude stream ecosystem: Potential response to climate change. Ecol. Appl. 2006, 16, 612–621. [Google Scholar] [CrossRef]
  141. Alvarez, D.; Nicieza, A.G. Compensatory response “defends” energy levels but not growth trajectories in brown trout, Salmo trutta L. Proc. R. Soc. B Biol. Sci. 2005, 272, 601–607. [Google Scholar] [CrossRef] [PubMed]
  142. Yvon-Durocher, G.; Jones, J.I.; Trimmer, M.; Woodward, G.; Montoya, J.M. Warming alters the metabolic balance of ecosystems. Philos. Trans. R. Soc. B Biol. Sci. 2010, 365, 2117–2126. [Google Scholar] [CrossRef] [PubMed][Green Version]
  143. McKee, D.; Atkinson, D.; Collings, S.; Eaton, J.; Harvey, I.; Heyes, T.; Hatton, K.; Wilson, D.; Moss, B. Macro-zooplankter responses to simulated climate warming in experimental freshwater microcosms. Freshw. Biol. 2002, 47, 1557–1570. [Google Scholar] [CrossRef]
  144. Mahdy, A.; Hilt, S.; Filiz, N.; Beklioğlu, M.; Hejzlar, J.; Özkundakci, D.; Papastergiadou, E.; Scharfenberger, U.; Šorf, M.; Stefanidis, K.; et al. Effects of water temperature on summer periphyton biomass in shallow lakes: A pan-European mesocosm experiment. Aquat. Sci. 2015, 77, 499–510. [Google Scholar] [CrossRef]
  145. Velthuis, M.; Domis, L.N.; Frenken, T.; Stephan, S.; Kazanjian, G.; Aben, R.; Hilt, S.; Kosten, S.; van Donk, E.; Van de Waal, D.B. Warming advances top-down control and reduces producer biomass in a freshwater plankton community. Ecosphere 2017, 8, e01651. [Google Scholar] [CrossRef][Green Version]
  146. Kazanjian, G.; Velthuis, M.; Aben, R.; Stephan, S.; Peeters, E.T.H.M.; Frenken, T.; Touwen, J.; Xue, F.; Kosten, S.; Van de Waal, D.B.; et al. Impacts of warming on top-down and bottom-up controls of periphyton production. Sci. Rep. 2018, 8. [Google Scholar] [CrossRef] [PubMed]
  147. Paerl, H.W.; Huisman, J. Climate change: A catalyst for global expansion of harmful cyanobacterial blooms. Environ. Microbiol. Rep. 2009, 1, 27–37. [Google Scholar] [CrossRef] [PubMed]
  148. Huber, V.; Wagner, C.; Gerten, D.; Adrian, R. To bloom or not to bloom: Contrasting responses of cyanobacteria to recent heat waves explained by critical thresholds of abiotic drivers. Oecologia 2012, 169, 245–256. [Google Scholar] [CrossRef] [PubMed]
  149. Kosten, S.; Huszar, V.L.M.; Bécares, E.; Costa, L.S.; van Donk, E.; Hansson, L.-A.; Jeppesen, E.; Kruk, C.; Lacerot, G.; Mazzeo, N.; et al. Warmer climates boost cyanobacterial dominance in shallow lakes. Glob. Chang. Biol. 2012, 18, 118–126. [Google Scholar] [CrossRef]
  150. Jensen, H.S.; Andersen, F.O. Importance of temperature, nitrate, and pH for phosphate release from aerobic sediments of four shallow, eutrophic lakes. Limnol. Oceanogr. 1992, 37, 577–589. [Google Scholar] [CrossRef][Green Version]
  151. Bannick, C.; Engelmann, B.; Fendler, J.; Frauenstein, J.; Ginzky, H.; Hornemann, C.; Ilvonen, O.; Kirschbaum, B.; Penn-Bressel, G.; Rechenberg, J.; et al. Groundwater in Germany; Federal Ministry for the Environment, Nature Conservation, Building and Nuclear Safety: Bonn, Germany, 2008; p. 72. [Google Scholar]
  152. Schuetz, T.; Weiler, M. Quantification of localized groundwater inflow into streams using ground-based infrared thermography. Geophys. Res. Lett. 2011, 38, L03401. [Google Scholar] [CrossRef]
  153. Sebok, E.; Duque, C.; Kazmierczak, J.; Engesgaard, P.; Nilsson, B.; Karan, S.; Frandsen, M. High-resolution distributed temperature sensing to detect seasonal groundwater discharge into Lake Vaeng, Denmark. Water Resour. Res. 2013, 49, 5355–5368. [Google Scholar] [CrossRef][Green Version]
  154. Whitehead, P.G.; Wilby, R.L.; Battarbee, R.W.; Kernan, M.; Wade, A.J. A review of the potential impacts of climate change on surface water quality. Hydrol. Sci. J. 2009, 54, 101–123. [Google Scholar] [CrossRef][Green Version]
  155. Vanek, V. Riparian Zone as a Source of Phosphorus for a Groundwater-Dominated Lake. Water Res. 1991, 25, 409–418. [Google Scholar] [CrossRef]
  156. Meinikmann, K.; Hupfer, M.; Lewandowski, J. Phosphorus in groundwater discharge—A potential source for lake eutrophication. J. Hydrol. 2015, 524, 214–226. [Google Scholar] [CrossRef]
  157. Ouyang, Y. Estimation of shallow groundwater discharge and nutrient load into a river. Ecol. Eng. 2012, 38, 101–104. [Google Scholar] [CrossRef]
  158. Perillon, C.; Poeschke, F.; Lewandowski, J.; Hupfer, M.; Hilt, S. Stimulation of epiphyton growth by lacustrine groundwater discharge to an oligo-mesotrophic hard-water lake. Freshw. Sci. 2017, 36, 555–570. [Google Scholar] [CrossRef]
  159. Phillips, G.; Willby, N.; Moss, B. Submerged macrophyte decline in shallow lakes: What have we learnt in the last forty years? Aquat. Bot. 2016, 135, 37–45. [Google Scholar] [CrossRef]
  160. Scheffer, M.; Hosper, S.; Meijer, M.; Moss, B.; Jeppesen, E. Alternative Equilibria in Shallow Lakes. Trends Ecol. Evol. 1993, 8, 275–279. [Google Scholar] [CrossRef]
  161. Hilt, S.; Koehler, J.; Adrian, R.; Monaghan, M.T.; Sayer, C.D. Clear, crashing, turbid and back—Long-term changes in macrophyte assemblages in a shallow lake. Freshw. Biol. 2013, 58, 2027–2036. [Google Scholar] [CrossRef]
  162. Hilt, S. Regime shifts between macrophytes and phytoplankton—Concepts beyond shallow lakes, unravelling stabilizing mechanisms and practical consequences. Limnetica 2015, 34, 467–479. [Google Scholar]
  163. Bradley, P.M.; Barber, L.B.; Duris, J.W.; Foreman, W.T.; Furlong, E.T.; Hubbard, L.E.; Hutchinson, K.J.; Keefe, S.H.; Kolpin, D.W. Riverbank filtration potential of pharmaceuticals in a wastewater-impacted stream. Environ. Pollut. 2014, 193, 173–180. [Google Scholar] [CrossRef] [PubMed]
  164. Haack, S.K.; Metge, D.W.; Fogarty, L.R.; Meyer, M.T.; Barber, L.B.; Harvey, R.W.; LeBlanc, D.R.; Kolpin, D.W. Effects on groundwater microbial communities of an engineered 30-Day in situ exposure to the antibiotic sulfamethoxazole. Environ. Sci. Technol. 2012, 46, 7478–7486. [Google Scholar] [CrossRef] [PubMed]
  165. Zhao, J.-L.; Zhang, Q.-Q.; Chen, F.; Wang, L.; Ying, G.-G.; Liu, Y.-S.; Yang, B.; Zhou, L.-J.; Liu, S.; Su, H.-C.; et al. Evaluation of triclosan and triclocarban at river basin scale using monitoring and modeling tools: Implications for controlling of urban domestic sewage discharge. Water Res. 2013, 47, 395–405. [Google Scholar] [CrossRef] [PubMed][Green Version]
  166. Eerkes-Medrano, D.; Thompson, R.C.; Aldridge, D.C. Microplastics in freshwater systems: A review of the emerging threats, identification of knowledge gaps and prioritisation of research needs. Water Res. 2015, 75, 63–82. [Google Scholar] [CrossRef] [PubMed]
  167. Conkle, J.L.; Gan, J.; Anderson, M.A. Degradation and sorption of commonly detected PPCPs in wetland sediments under aerobic and anaerobic conditions. J. Soils Sediments 2012, 12, 1164–1173. [Google Scholar] [CrossRef]
  168. Ebele, A.J.; Abou-Elwafa Abdallah, M.; Harrad, S. Pharmaceuticals and personal care products (PPCPs) in the freshwater aquatic environment. Emerg. Contam. 2017, 3, 1–16. [Google Scholar] [CrossRef]
  169. Drury, B.; Rosi-Marshall, E.; Kelly, J.J. Wastewater treatment effluent reduces the abundance and diversity of benthic bacterial communities in urban and suburban rivers. Appl. Environ. Microbiol. 2013, 79, 1897–1905. [Google Scholar] [CrossRef] [PubMed]
  170. Roy, J.W.; Malenica, A. Nutrients and toxic contaminants in shallow groundwater along Lake Simcoe urban shorelines. Inland Waters 2013, 3, 125–138. [Google Scholar] [CrossRef]
  171. Sand-Jensen, K. Environmental variables and their effect on photosynthesis of aquatic plant communities. Aquat. Bot. 1989, 34, 5–25. [Google Scholar] [CrossRef]
  172. Maberly, S.C.; Madsen, T.V. Affinity for CO2 in relation to the ability of freshwater macrophytes to use HCO−3. Funct. Ecol. 1998, 12, 99–106. [Google Scholar] [CrossRef]
  173. Demars, B.O.L.; Tremolieres, M. Aquatic macrophytes as bioindicators of carbon dioxide in groundwater fed rivers. Sci. Total Environ. 2009, 407, 4752–4763. [Google Scholar] [CrossRef] [PubMed]
  174. Cole, J.J.; Caraco, N.F.; Kling, G.W.; Kratz, T.K. Carbon Dioxide Supersaturation in the Surface Waters of Lakes. Science 1994, 265, 1568–1570. [Google Scholar] [CrossRef] [PubMed]
  175. Weyhenmeyer, G.A.; Kosten, S.; Wallin, M.B.; Tranvik, L.J.; Jeppesen, E.; Roland, F. Significant fraction of CO2 emissions from boreal lakes derived from hydrologic inorganic carbon inputs. Nat. Geosci. 2015, 8, 933–936. [Google Scholar] [CrossRef]
  176. Marce, R.; Obrador, B.; Morgui, J.-A.; Lluis Riera, J.; Lopez, P.; Armengol, J. Carbonate weathering as a driver of CO2 supersaturation in lakes. Nat. Geosci. 2015, 8, 107–111. [Google Scholar] [CrossRef]
  177. Sand-Jensen, K.; Borum, J. Interactions among phytoplankton, periphyton, and macrophytes in temperate freshwaters and estuaries. Aquat. Bot. 1991, 41, 137–175. [Google Scholar] [CrossRef]
  178. Sand-Jensen, K.; Staehr, P.A. CO2 dynamics along Danish lowland streams: Water-air gradients, piston velocities and evasion rates. Biogeochemistry 2012, 111, 615–628. [Google Scholar] [CrossRef]
  179. Macpherson, G.L. CO2 distribution in groundwater and the impact of groundwater extraction on the global C cycle. Chem. Geol. 2009, 264, 328–336. [Google Scholar] [CrossRef]
  180. Vesper, D.J.; Edenborn, H.M. Determination of free CO2 in emergent groundwaters using a commercial beverage carbonation meter. J. Hydrol. 2012, 438–439, 148–155. [Google Scholar] [CrossRef]
  181. Hutchinson, G.E. A Treatise on Limnology. 3: Limnological Botany; Wiley: New York, NY, USA, 1975; ISBN 978-0-471-42574-8. [Google Scholar]
  182. Moss, B. Ecology of Fresh Waters: Man and Medium, 2nd ed.; Blackwell Science: Oxford, UK, 1996; ISBN 978-0-632-01642-6. [Google Scholar]
  183. Jones, J.I. The metabolic cost of bicarbonate use in the submerged plant Elodea nuttallii. Aquat. Bot. 2005, 83, 71–81. [Google Scholar] [CrossRef]
  184. Olesen, B.; Madsen, T.V. Growth and physiological acclimation to temperature and inorganic carbon availability by two submerged aquatic macrophyte species, Callitriche cophocarpa and Elodea canadensis. Funct. Ecol. 2000, 14, 252–260. [Google Scholar] [CrossRef]
  185. Maberly, S.C. Photosynthesis by Fontinalis antipyretica. I. New Phytol. 1985, 100, 127–140. [Google Scholar] [CrossRef]
  186. Maberly, S.C. Photosynthesis by Fontinalis antipyretica. II. New Phytol. 1985, 100, 141–155. [Google Scholar] [CrossRef]
  187. Maberly, S.C.; Berthelot, S.A.; Stott, A.W.; Gontero, B. Adaptation by macrophytes to inorganic carbon down a river with naturally variable concentrations of CO2. J. Plant Physiol. 2015, 172, 120–127. [Google Scholar] [CrossRef] [PubMed]
  188. Frandsen, M.; Nilsson, B.; Engesgaard, P.; Pedersen, O. Groundwater seepage stimulates the growth of aquatic macrophytes. Freshw. Biol. 2012, 57, 907–921. [Google Scholar] [CrossRef]
  189. Takahashi, K.; Asaeda, T. The effect of spring water on the growth of a submerged macrophyte Egeria densa. Landsc. Ecol. Eng. 2012, 10, 99–107. [Google Scholar] [CrossRef][Green Version]
  190. Korner, S. Development of submerged macrophytes in shallow Lake Muggelsee (Berlin, Germany) before and after its switch to the phytoplankton-dominated state. Arch. Hydrobiol. 2001, 152, 395–409. [Google Scholar] [CrossRef]
  191. Maberly, S.C. The fitness of the environments of air and water for photosynthesis, growth, reproduction and dispersal of photoautotrophs: An evolutionary and biogeochemical perspective. Aquat. Bot. 2014, 118, 4–13. [Google Scholar] [CrossRef]
  192. Gooddy, D.C.; Shand, P.; Kinniburgh, D.G.; Van Riemsdijk, W.H. Field-based partition coefficients for trace elements in soil solutions. Eur. J. Soil Sci. 1995, 46, 265–285. [Google Scholar] [CrossRef]
  193. Monteith, D.T.; Stoddard, J.L.; Evans, C.D.; de Wit, H.A.; Forsius, M.; Hogasen, T.; Wilander, A.; Skjelkvale, B.L.; Jeffries, D.S.; Vuorenmaa, J.; et al. Dissolved organic carbon trends resulting from changes in atmospheric deposition chemistry. Nature 2007, 450, 537–540. [Google Scholar] [CrossRef] [PubMed]
  194. Reckhow, D.; Singer, P.; Malcolm, R. Chlorination of humic materials: Byproduct formation and chemical interpretations. Environ. Sci. Technol. 1990, 24, 1655–1664. [Google Scholar] [CrossRef]
  195. Reyjol, Y.; Argillier, C.; Bonne, W.; Borja, A.; Buijse, A.D.; Cardoso, A.C.; Daufresne, M.; Kernan, M.; Ferreira, M.T.; Poikane, S.; et al. Assessing the ecological status in the context of the European Water Framework Directive: Where do we go now? Sci. Total Environ. 2014, 497, 332–344. [Google Scholar] [CrossRef] [PubMed]
  196. Birk, S.; Bonne, W.; Borja, A.; Brucet, S.; Courrat, A.; Poikane, S.; Solimini, A.; van de Bund, W.; Zampoukas, N.; Hering, D. Three hundred ways to assess Europe’s surface waters: An almost complete overview of biological methods to implement the Water Framework Directive. Ecol. Indic. 2012, 18, 31–41. [Google Scholar] [CrossRef]
  197. Jeppesen, E.; Jensen, J.P.; Sondergaard, M.; Lauridsen, T.; Landkildehus, F. Trophic structure, species richness and biodiversity in Danish lakes: Changes along a phosphorus gradient. Freshw. Biol. 2000, 45, 201–218. [Google Scholar] [CrossRef]
  198. Kuiper, J.J.; Verhofstad, M.J.J.M.; Louwers, E.L.M.; Bakker, E.S.; Brederveld, R.J.; van Gerven, L.P.A.; Janssen, A.B.G.; de Klein, J.J.M.; Mooij, W.M. Mowing submerged macrophytes in shallow lakes with alternative stable states: Battling the good guys? Environ. Manag. 2017, 59, 619–634. [Google Scholar] [CrossRef] [PubMed]
  199. Hilt, S.; Henschke, I.; Ruecker, J.; Nixdorf, B. Can submerged macrophytes influence turbidity and trophic state in deep lakes? Suggestions from a case study. J. Environ. Qual. 2010, 39, 725–733. [Google Scholar] [CrossRef] [PubMed]
  200. Rastogi, R.P.; Sinha, R.P.; Incharoensakdi, A. The cyanotoxin-microcystins: Current overview. Rev. Environ. Sci. Biol. Technol. 2014, 13, 215–249. [Google Scholar] [CrossRef]
  201. Chorus, I.; Bartram, J. (Eds.) Toxic Cyanobacteria in Water: A Guide to Their Public Health Consequences, Monitoring, and Management; E & FN Spon: London, UK; New York, NY, USA, 1999; ISBN 978-0-419-23930-7. [Google Scholar]
  202. Heisler, J.; Glibert, P.M.; Burkholder, J.M.; Anderson, D.M.; Cochlan, W.; Dennison, W.C.; Dortch, Q.; Gobler, C.J.; Heil, C.A.; Humphries, E.; et al. Eutrophication and harmful algal blooms: A scientific consensus. Harmful Algae 2008, 8, 3–13. [Google Scholar] [CrossRef] [PubMed][Green Version]
  203. Grützmacher, G.; Wiese, B.; Hülsoff, I.; Orlikowski, D.; Hoa, E.; Moreau-Le Golvan, Y. Bank Filtration and Aquifer Recharge for Drinking Water Production: Application, Efficiency and Perspectives—An Integration of NASRI Outcomes and International Experiences; Kompetenzzentrum Wasser Berlin gGmbH: Berlin, Germany, 2011. [Google Scholar]
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Figure 1. During induced bank filtration surface water from a source water body (dark blue) infiltrates into the sub-surface and reaches the groundwater well (light blue with dots). Traditionally, research focused on the effects of surface water and bank filtration on drinking water quality and quantity (unfilled arrows). We propose to include the neglected effects of bank filtration on surface water quality (filled arrow) into research to secure sustainable drinking water supply and sufficient ecological quality of source water bodies.
Figure 1. During induced bank filtration surface water from a source water body (dark blue) infiltrates into the sub-surface and reaches the groundwater well (light blue with dots). Traditionally, research focused on the effects of surface water and bank filtration on drinking water quality and quantity (unfilled arrows). We propose to include the neglected effects of bank filtration on surface water quality (filled arrow) into research to secure sustainable drinking water supply and sufficient ecological quality of source water bodies.
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Figure 2. Top 10 studies in the categories research area (a) and countries/regions (b) of papers available on bank filtration (Web of Science topic search “*bank filtration”, June 2018).
Figure 2. Top 10 studies in the categories research area (a) and countries/regions (b) of papers available on bank filtration (Web of Science topic search “*bank filtration”, June 2018).
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Figure 3. Bank filtration as part of a semi-closed water cycle. The water bodies of Berlin with waterworks, wastewater treatment plants and IBF abstraction well galleries, after Berlin Water Utilities [76] and Jekel et al. [77]. Most of the Berlin surface waters are part of a semi-closed water cycle where water is being abstracted via bank filtration where treated wastewater is released.
Figure 3. Bank filtration as part of a semi-closed water cycle. The water bodies of Berlin with waterworks, wastewater treatment plants and IBF abstraction well galleries, after Berlin Water Utilities [76] and Jekel et al. [77]. Most of the Berlin surface waters are part of a semi-closed water cycle where water is being abstracted via bank filtration where treated wastewater is released.
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Figure 4. Bank filtration at Lake Müggelsee, Berlin, Germany. (a) Groundwater isolines for 2015 (purple lines with numbers indicate groundwater levels in meter above sea-level), well galleries with a total of 170 groundwater abstraction wells (red lines), groundwater monitoring wells (green points) around Lake Müggelsee (Germany). The transect of groundwater levels shown in panel b is indicated with dashed black line. Background map including wells provided by the Senate of Berlin, Department of urban development and housing via the FISBroker (Fachübergreifendes InformationsSystem) online mapping tool [81]. (b) Groundwater (GW) levels (MASL = meter above sea lelvel) from 1905 (before groundwater abstraction started) and 1990 (with active pumping) along a north-south transect north of Lake Müggelsee (after Frey et al. [82]). The transect is indicated in panel a. (c) Yearly abstraction rates (solid line) of all wells surrounding Lake Müggelsee (galleries A–F) 1959–2014 [80] and surface water discharge (dotted line) into Lake Müggelsee 1976–2014 [83,84].
Figure 4. Bank filtration at Lake Müggelsee, Berlin, Germany. (a) Groundwater isolines for 2015 (purple lines with numbers indicate groundwater levels in meter above sea-level), well galleries with a total of 170 groundwater abstraction wells (red lines), groundwater monitoring wells (green points) around Lake Müggelsee (Germany). The transect of groundwater levels shown in panel b is indicated with dashed black line. Background map including wells provided by the Senate of Berlin, Department of urban development and housing via the FISBroker (Fachübergreifendes InformationsSystem) online mapping tool [81]. (b) Groundwater (GW) levels (MASL = meter above sea lelvel) from 1905 (before groundwater abstraction started) and 1990 (with active pumping) along a north-south transect north of Lake Müggelsee (after Frey et al. [82]). The transect is indicated in panel a. (c) Yearly abstraction rates (solid line) of all wells surrounding Lake Müggelsee (galleries A–F) 1959–2014 [80] and surface water discharge (dotted line) into Lake Müggelsee 1976–2014 [83,84].
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Figure 5. Conceptual scheme of potential links between induced bank filtration and relevant parameters and processes concerning surface water quality (filled arrows, DIC = dissolved inorganic carbon, DOC = dissolved organic carbon). Changes in surface water quality will in turn affect the quality of bank filtrate, which has been the focus of previous research (unfilled arrows).
Figure 5. Conceptual scheme of potential links between induced bank filtration and relevant parameters and processes concerning surface water quality (filled arrows, DIC = dissolved inorganic carbon, DOC = dissolved organic carbon). Changes in surface water quality will in turn affect the quality of bank filtrate, which has been the focus of previous research (unfilled arrows).
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Table
Table 1. Percentage of induced bank filtration (IBF) in drinking water supply of different countries/cities and source surface water bodies. The symbol ” means same as above, * indicate that info is missing, - that info is not applicable.
Table 1. Percentage of induced bank filtration (IBF) in drinking water supply of different countries/cities and source surface water bodies. The symbol ” means same as above, * indicate that info is missing, - that info is not applicable.
CountryCityPercentage of Drinking Water Provided by IBFSource Water BodiesRiver Discharge, Lake Volume/SizeReference
Austria**River Enns65–206 m3/s[17]
”Innsbruck*River Inn730 m3/s[18]
”Vienna, Linz*Danube1900 m3/s[19]
Bulgaria***-[19]
FinlandKuopio*Lake Kallavesi4730 million m3[20,21]
”**Not mentioned*[9]
FranceParis region*Seine River450 m3/s[22]
Germany-9 to 16--[23,24]
”Berlin60--[4]
””-Lake Müggelsee36 million m3[5]
””-Lake Tegel26 million m3[5,25,26,27,28,29,30,31,32]
””-Lake Wannsee15 million m3[5,12,27,33,34,35,36,37]
”Radeburg-Radeburg Reservoir0.35 km2
Max depth: 3 m		[38]
”Düsseldorf~100River Rhine2300 m3/s[7]
”Frankfurt am Main*River Rhine2300 m3/s[39]
””*Lower River Main193 m3/s[39]
”Torgau and Mockritz*Elbe River700 m3/s[40]
Hungary-45--[7]
”Budapest-Danube6460 m3/s[40]
”*-Rivers Raba, Drava, Ipoly, Sajo, Hernád17, 500,
21, 67, 27 m3/s		[41]
ItalyLucca, Pisa, Livorno(300,000 inhabitants)River Serchio46 m3/s[42]
LatviaRiga*Lake Mazais
Baltezers
Lake Lielais Baltezers		10 million m3
18 million m3		[43]
”
The Netherlands-5--[7]
”Remmerden*River Rhine2300 m3/s[44]
”Zwijndrecht*River Rhine2300 m3/s”
”Roosteren*River Meuse276 m3/s”
”Roermond*Gravel pit lake De Lange Vlieter1.2 km2
Max depth: 35 m		[45,46,47]
Norway
”		Hemne
”		*
*		Lake Rovatnet
River Buga		8 km2
*		[48]
PolandPoznań*River Warta60 m3/s[49]
RomaniaIasi*Moldova River143 m3/s[50]
Slovak Republic-50--[7]
SloveniaMaribor-Drava River500 m3/s[40]
Switzerland-
 		10–30-
River Thur		-
40–50 m3/s		[13,51]
UK**Streams Wissey, Rhee and Pang1.9, 1.25, 0.64 m3/s[52]
Canada**Lake A and B (artificial)*[10]
USAJeffersonville*Ohio River3512 m3/s[53,54]
”Santa Rosa*Russian River66 m3/s[55]
”Cincinnati*Great Miami River109 m3/s[56]
”Columbus*Scioto/Big Walnut Creek6 m3/s”
”Galesburg*Mississippi River16,792 m3/s”
”Independence
Kansas City Parkville		*
 
Missouri River2158 m3/s”
”
[54]
”Jacksonville*Illinois River659 m3/s[56]
”Kalama*Kalama River40 m3/s”
”Kennewick*Columbia River7500 m3/s”
”Lincoln*Platte River203 m3/s”
”Mt. Carmel Terre Haute*
 		Wabash River837 m3/s”
[54]
”Sacramento*Sacramento River660 m3/s”
”Cape Cod*Ashumet Pond6 million m3[57]
Brazil**Beberibe River0.3–0.4 m water depth[58]
”**Lake Lagoa do Peri36 million m3[59,60]
ChinaMatan96Yellow River1839 m3/s[61]
”Baisha Town82.1Yangtze River31,100 m3/s”
”Jiuwutan82.6Yellow River1839 m3/s”
”Qingpu district70–80Taipu River300 m3/s”
”Xuzhou>80Kui River*”
”Chengdu80Yinma River30 m3/s”
India-*--[62]
”Delhi*Yamuna River100–1300 m3/s[63]
”Satpuli
Srinagar		*
*		East Nayar River
River Alaknanda		-
507 m3/s		[64]
”Haridwar*River Ganga1455 m3/s[65]
”Nainital*Lake Nainital6 million m3[66]
MalaysiaKuala Kangsar*Sungai Perak (river)57 m3/s[67]
South Korea**Nakdong River37–3462 m3/s[68]
ThailandChiang Mai*Ping River287 m3/s[69]
Egypt-0.1 (increasing)Upper Nile1548 m3/s[70]
”Sidfa*Nile2830 m3/s[71]
”Aswan*””[72]
Table
Table 2. Potential effects of induced bank filtration on physical (WT = water temperature, RT = retention time, WL = water level) and chemical (DIC = dissolved inorganic carbon, DOC = dissolved organic carbon, pollutants such as pharmaceutical remnants and microplastics) parameters in surface waters (mechanisms: D: directly, I: indirectly due to the interruption of groundwater discharge) and examples for subsequent effects on biological parameters. The symbol “+” means that the effect will enhance the affected biological parameter, “−” that the effect will decrease the parameter, and “?” that the outcome is uncertain. The predicted effects within the parameter categories are roughly ordered according to our estimate of likelihood of occurring.
Table 2. Potential effects of induced bank filtration on physical (WT = water temperature, RT = retention time, WL = water level) and chemical (DIC = dissolved inorganic carbon, DOC = dissolved organic carbon, pollutants such as pharmaceutical remnants and microplastics) parameters in surface waters (mechanisms: D: directly, I: indirectly due to the interruption of groundwater discharge) and examples for subsequent effects on biological parameters. The symbol “+” means that the effect will enhance the affected biological parameter, “−” that the effect will decrease the parameter, and “?” that the outcome is uncertain. The predicted effects within the parameter categories are roughly ordered according to our estimate of likelihood of occurring.
ParameterPredicted EffectMechanismReferencesAffected Biological ParameterEffectReferences (Example)
PhysicalHigher summer WTI[93]Biodiversity
Macrophyte dominance
Harmful blooms		±
−
+		[94]
[95]
[95]
Lower winter WTI[96]Biodiversity
Macrophyte dominance
Harmful blooms		?
−
?		[97]
 
 
Higher RTD, I Biodiversity
Macrophyte presence
Harmful blooms		±
±
±		[98]
[92]
[99]
Lower flowD, I[89]Biodiversity
Macrophyte presence
Harmful blooms		±
±
+		[87]
[100]
[101]
Sediment cloggingD
 		[26]Biodiversity
Macrophyte dominance
Harmful blooms		?
±
?		[102]
 
 
Lower WLD, I[103]Biodiversity
Macrophyte presence
Harmful blooms		− (?)
−
+		[103]
[99]
 
Stronger WL fluctuationsD, I[103]Biodiversity
Macrophyte presence
Harmful blooms		±
−
?		[104]
[105]
 
ChemicalLower DICI[106]Biodiversity
Macrophyte dominance
Harmful blooms		±
−
±		[98]
[107]
[108]
Lower external nutrient loadI Biodiversity
Macrophyte dominance
Harmful blooms		+
+
−		[109]
[110]
[108]
Lower DOCI[106]Biodiversity
Macrophyte dominance
Harmful blooms		+ (?)
+ (?)
+/−		[111]
[112]
[113]/[114]
Lower pollutant loadI Biodiversity
Macrophyte dominance
Harmful blooms		+
+ (?)
?		[115]
 
 
Higher pollutant load in the littoralI Biodiversity
Macrophyte dominance
Harmful blooms		−
− (?)
?		[115]
 
 

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MDPI and ACS Style

Gillefalk, M.; Massmann, G.; Nützmann, G.; Hilt, S. Potential Impacts of Induced Bank Filtration on Surface Water Quality: A Conceptual Framework for Future Research. Water 2018, 10, 1240. https://doi.org/10.3390/w10091240

AMA Style

Gillefalk M, Massmann G, Nützmann G, Hilt S. Potential Impacts of Induced Bank Filtration on Surface Water Quality: A Conceptual Framework for Future Research. Water. 2018; 10(9):1240. https://doi.org/10.3390/w10091240

Chicago/Turabian Style

Gillefalk, Mikael, Gudrun Massmann, Gunnar Nützmann, and Sabine Hilt. 2018. "Potential Impacts of Induced Bank Filtration on Surface Water Quality: A Conceptual Framework for Future Research" Water 10, no. 9: 1240. https://doi.org/10.3390/w10091240

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