Phosphorous Supply to a Eutrophic Artiﬁcial Lake: Sedimentary versus Groundwater Sources

: The eutrophic Lake Eichbaumsee, a ~1 km long and 280 m wide (maximum water depth 16 m) dredging lake southeast of Hamburg (Germany), has been treated for water quality im-provements using various techniques (i.e., aeration plants, removal of dissolved phosphorous by aluminum phosphorous precipitation, and by Bentophos ® (Phoslock Environmental Technologies, Sydney, Australia), adsorption) during the past ~15 years. Despite these treatments, no long-term improvement of the water quality has been observed and the lake water phosphorous content has continued to increase by e.g., ~670 kg phosphorous between autumn 2014 and autumn 2019. As no creeks or rivers drain into the lake and hydrological groundwater models do not suggest any major groundwater discharge into the lake, sources of phosphorous (and other nutrients) are unknown. We investigated the phosphorous ﬂuxes from sediment pore water and from groundwater in the water body of the lake. Sediment pore water was extracted from sediment cores recovered by divers in August 2018 and February 2019. Diffusive phosphorous ﬂuxes from pore water were calculated based on phosphorus gradients. Stable water isotopes ( δ 2 H, δ 18 O) were measured in the lake water, in interstitial waters in the banks surrounding the lake, in the Elbe River, and in three groundwater wells close to the lake. Stable isotope ( δ 2 H, δ 18 O) water mass balance models were used to compute water inﬂow/outﬂow to/from the lake. Our results revealed pore-water borne phosphorous ﬂuxes between 0.2 mg/m 2 /d and 1.9 mg/m 2 /d. Assuming that the measured phosphorous ﬂuxes are temporarily and spatially representative for the whole lake, about 11 kg/a to 110 kg/a of phosphorous is released from sediments. This amount is lower than the observed lake water phosphorous increase of ~344 kg between April 2018 and November 2018. Water stable isotope ( δ 2 H, δ 18 O) compositions indicate a water exchange between an aquifer and the lake water. Based on stable isotope mass balances we estimated an inﬂow of phosphorous from the aquifer to the lake of between ~150 kg/a and ~390 kg/a. This result suggests that groundwater-borne phosphorous is a signiﬁcant phosphorous source for the Eichbaumsee and highlights the importance of groundwater for lake water phosphorous balances.


Introduction
Eutrophication is one of the main reasons for the poor ecological status of many lakes around the world, causing the deterioration of lake water quality and occurrences of harmful cyanobacteria blooms [1,2]. High nutrient (i.e., phosphorus and nitrogen) supply via river and groundwater discharge or (to a minor degree) by atmospheric deposition are body ( Figure 2). Further lake water treatments using Bentophos ® (a lanthanum-modified bentonite; Bentophos ® ) were applied, which are expected to bind and fix phosphorous permanently in the sediments. Each November between 2010 and 2013 Bentophos ® was added to the lake (288 t in total) [28]. The Bentophos ® treatments in November 2010 and 2013 removed about 90% of the phosphorous from the water column ( Figure 2). After the last Bentophos ® application it was decided by the local authorities that no further treatments would be necessary [28]. However, in 2014 phosphorous increased by 306 kg between spring and autumn with little reduction during full water column circulation in the following winter 2015 (Figure 2). Since then, lake water phosphorus concentrations and eutrophication levels remained high triggering harmful algal blooms regularly during the spring and summer [29]. The application of Bentophos ® in the winters 2011 through 2013 did not lead to a long-term reduction of the lake phosphorous content (modified from [28]).
These unsuccessful remediation attempts have raised the question as to whether the lake phosphorous increase is related to the failure of Bentophos ® to fix phosphorous in the sediments permanently or if an additional and so far unconsidered external phosphorous source exists. As no creeks or rivers discharge into the lake and available hydrological groundwater models do not suggest any major groundwater flow into the lake, the nature of any such external phosphorous source remained unknown.
In our study we investigated the internal and external phosphorous supply to Lake Eichbaumsee using sediment geochemistry and sediment-porewater flux measurements to quantify the internal phosphorous source. Based on water stable isotope mass balances we discuss a groundwater-borne phosphorous supply. To our knowledge this is the first study combing such methodological approaches to constrain a lake phosphorous balance. Our results reveal that, apart from the internal phosphorous cycling, groundwaterborne phosphorous is a very important source for the overall phosphorous budget of the lake maintaining its eutrophication level. This result highlights the importance of recognizing groundwater as a phosphorous source before lake water remediation measures are designed.

Study Area
Lake Eichbaumsee is located southeast of Hamburg (Figure 1) in the North German sedimentary plain. During several Quaternary glaciations/deglaciation periods meltwaters formed the River Elbe valley and deposited a wide range of glacial sediments (sands, tills, gravels). Below a~5-10 m thick clayey layer~25-30 m thick Holocene sands were deposited. Extractions of these sands for road constructions in 1972-1976 formed the Lake Eichbaumsee. The artificial lake is max. 1 km long and 280 m wide with an average depth of 6.5 m and a maximum depth of 16 m. The lake s surface area is 240,000 m 2 and the water volume is around 1,727,000 m 3 [28]. The lake used to be directly connected to a branch of River Elbe, the so-called Dove-Elbe, until its straightening in 1977. No creeks drain into the lake. To the north of the lake an east-west trending dyke (Moorfleeter Deich, Figure 1) was constructed in the 19th century to protect the hinterland with farming land from possible flooding of the Elbe River. Two natural reserves border the Eichbaumsee to the east and to the south. The lake has an important function as a local recreation area and as the lake is used for swimming, water quality has been monitored regularly throughout the years by local authorities.

Sampling Locations
Three sampling campaigns were carried out: 10-11 April 2018, 14-15 August 2018, and 26-27 February 2019. During these campaigns, sediments, sediment pore waters, ground water samples, lake water column samples, as well as pore water samples from the lake banks were collected. An overview of the sampling and sampling locations is given in Figure 1 and Table 1. Three deep locations in the lake (E2: 14 m, E3: 13 m, and E: 16 m) were selected as coring stations because the highest accumulation of soft sediments was expected here, a situation that is favorable for sediment core recovery by divers who pushed PVC liners (with lengths between 0.5-1 m, diameter 10 cm) into the soft lake sediment. Upon recovery, sediment pore water was collected from the cores by pushing Rhizons into pre-drilled holes in the liner (intervals between 1 and 2 cm) and extracting pore water following the procedure given in Seeberg-Elverfeldt et al. [30]. For analysis of major and minor ions one fraction of pore water was immediately acidified with HNO 3 in order to prevent the formation of hydrous ferric oxides and potential co-precipitation/sorption. A second fraction was stored in a freezer for later nutrient analyses. Furthermore, 2 mL of pore water was sampled in glass vials for water stable isotope analyses (D/O). Due to a failure in the preservation of pore water samples during the April campaign, we only report pore water data from August and February. In April, after pore-water sampling, sediments were extruded from the core and sampled in 1-5 cm thick slices for geochemical analyses.

Push-Point Sampling
From the lake banks pore water samples were obtained from 0.5-1.8 m depths using a push-point sampler [31]. The water samples were 0.45 µm filtered and separated in aliquots similar to the sediment pore water samples (see Section 2.2.1).

Lake Water and Groundwater Sampling
Lake water sampling for the chemical and physical characterization of the lake water body was conducted as full water-column profile sampling. Water temperature and oxygen content were determined using in situ sensors (EXO multiparameter sonde, YSI, Yellow Springs, OH, USA). Water samples were obtained in 0.3 m, 3 m, 6 m, 9 m, and 12 m water depths using Niskin ® bottles.
The area of Hamburg hosts one of the most productive groundwater reserves in Europe with six main aquifers found in Quaternary to Tertiary sediments. The deepest aquifer (~500 m depth below surface) is found in sandy layers (Braunkohlesande), whereas the shallowest aquifer sits 2-5 m below surface in sandy Weichsel sediments. Groundwater samples were taken from three wells located in the vicinity of the lake. The wells GW979 and GW5028 tap water from Weichsel sandy sediments, whereas GW11028 waters are from sandy-silty sediments of the Saale age. The respective groundwater levels sit between 0.8 m and 0.2 m above sea-level with seasonal fluctuations of less than 0.5 m (note that the Lake Eichbaumsee level is~0.5 m above sea-level). All groundwater samples were obtained after purging the wells for about 15 min until water conductivity remained constant. The samples were 0.45 µm filtered, and aliquots for major and minor element analyses, nutrients, and stable isotopes (D/O) were obtained following the procedure outlined above (see Section 2.2.1).

Water Samples
Phosphorous in water samples was analyzed as ortho-phosphate (PO 4 3− ) using continuous flow analysis (CFA) (Allianz Instruments). Phosphorus, iron, manganese, and lanthanum analyses in all water samples were performed using an Agilent 8800 ICP-MS/MS in mass-shift-mode (m/z 31 → m/z 47) after reaction with O 2 to avoid 14 N 16 OH +interferences. Accuracy was checked by analyzing phosphorous and iron standard solutions (Merck KGaA). The precisions ranged between ±10% (ICP-MS) and ±20% (CFA).

Major Elements in Sediments
For the analyses of major elements (Si, Al, Fe, Mn, Mg, Ca, Ki, Ti, P) about 200 mg of ground sediments were dissolved in a mixture of HNO 3 , HCL, and HF. Measurements were conducted in the presence of an yttrium internal standard using ICP-OES (Inductively Coupled Plasma Optical Emission Spectrometry, SPECTRO Ciros). For measurements of lanthanum (La) we used a Quadrupole ICP-MS (Agilent 7500 cc) with an Octopol reaction system. For the quality control of the measurements three sediment/rock standards (GSMS-2; BHVO_2, BIR_1a) were used. The analytical precision was ≤2%. In order to correct for the loss of silica during sample preparation, loss on ignition was determined by weighing sediment samples before and after heating to 1000 • C.

Stable Isotope Analyses
Water stable isotope analyses were carried out using laser cavity ring-down spectroscopy (Picarro L2120-i, Santa Clara, CA, USA). The results of the hydrogen and oxygen isotope measurements are expressed as delta notations (δ 2 H, δ 18 O) relative to the Vienna standard mean ocean water (VSMOW). The analytical uncertainties of δ 18 O and δ 2 H measurements are ± 0.1‰ and ± 0.8‰, respectively.

Calculation of Pore Water Fluxes
Diffusive phosphorous fluxes from the sediment pore water to the overlying lake water were calculated following first Fick's law [32] where F is diffusive flux, n is the sediment porosity (-), D s is the diffusion coefficient (cm 2 s −1 ), and ∂C ∂x is the concentration gradient (mg cm −2 ). The porosity was determined from the sediment water content; the diffusion coefficient was calculated following Ullman & Aller [33].

Water Stable Isotope Mass Balance
The determination of groundwater discharge into the lake based on a D/O mass balance assumes that the D/O signature of the lake water is controlled by (i) the physical lake water/groundwater interaction, (ii) precipitation, and (iii) evaporation [16,17], with δ P , δ Gi , δ E , and δ Go (‰) being the isotopic signatures of precipitation (P (mm)), groundwater discharge (G i (m 3 day −1 )), evaporation (E (mm)) and groundwater exfiltration (G o (m 3 day −1 )), respectively. A constant lake volume (V (m 3 )) over time is assumed. Furthermore, the isotopic signature of the lake water (δ L ) is assumed to be representative of the lake water that is exfiltrating into the aquifer (δ Go ) [16].
The seasonality factor accounts for the changes in the isotopic fractionation during evaporation in a seasonal cycle with k = 0.5 for strong seasonal climate alterations and k = 1 for climates with minor seasonality. Estimates of k are obtained by dual analysis of δ 18 O and δ 2 H. A fit of the mean annual evaporation flux-weighted δ A is adjusted by optimizing k to fit δ E to the local evaporation line [37,38].
The remaining unknown parameter that has to be calculated to solve Equation (2) is the amount of groundwater discharging into the lake (G i ) as given by Knöller et al. [17] (Equation (5)): with S i being surface water runoff (m 3 ). S i was assumed to be the part of the precipitation that falls on the surroundings of the lake (i.e., the slope area around the lake) and subsequently drains into the lake [17]. A footpath around the lake was considered as the highest area around the lake, and the area between this path and the lake was chosen as the slope area for the S i calculation.

Water Column
Water column profiles of temperature and oxygen were obtained at locations E2 and E4 between April 2018 and February 2019 (Figures 3 and 4); phosphorous was measured at E2 only ( Figure 5). Figure 3 illustrates that the lake water column is characterized by a well-developed thermocline during the summer months (May, July, and August) and a homogenous temperature distribution in the winter (November and February).   Whereas oxygen was present throughout the water column in April, the water column below 6 m water depth was anoxic in July and August ( Figure 4). In November, only the water column >10 m was still anoxic whereas in February oxygen was found throughout the water column. At location E4 with the greatest water depth (16 m) the water column >15 m was still anoxic in February.
Phosphorous concentrations in the water column were relatively constant ranging between 0.29 and 0.35 mg/L in April ( Figure 5). In May the deep water (12 m) concentration increased to 0.69 mg/L. In July and August, a steep gradient developed between high phosphorous concentrations in deep waters (July: 1.20 mg/L; August: 1.30 mg/L) and low concentrations at 0.3 m, i.e., shallow water depth (July: 0.09 mg/L; August 0.09 mg/L). Whereas these high phosphorous concentrations were still present in November in deep water (1.10 mg/L at 12 m), in water depths ≤9 m phosphorous was found to be in a relatively small range (0.47-0.51 mg/L). A homogeneous phosphorous distribution (range 0.41-0.44 mg/L) throughout the water column was observed in February 2019.
The water column distributions of iron and manganese are characterized by seasonal changes comparable to those of phosphorous. In spring (April 2018, February 2019) both elements showed relatively low and constant concentrations ( Figure 6). In May, July, and August, however, gradients between high deep-water and low mid-water concentrations were observed. Mixing of the water column caused homogenous distributions of iron and manganese in the water column at ≤9 m in November with elevated concentrations at a water depth of 12 m.

Phosphorous, Iron, and Lanthanum in Sediments
The obtained sediments consisted generally of homogeneous greyish-black silty mud. Sediments at the base of core E3 were composed of a light-greyish sand layer.
The results of the geochemical analyses of sediments can be found in the supplementary materials (Table S2).
The generally lower phosphorous, iron, manganese, and lanthanum concentrations in surface sediments at location E4 compared to E2 and E3 are probably related to higher SiO 2 concentrations in the E4 core causing a higher dilution of the other elements (Table S2).

Phosphorous Distribution in Sediment Pore Waters
As stated above, phosphorus was measured as ortho-phosphate via CFA and as total phosphorus via ICP-MS. The close relationship between ortho-phosphate (as P) and total phosphorus suggests that most of the phosphorus is bound as ortho-phosphate (Figure 8). At locations E2 and E4 the sediment pore water phosphorous concentrations in August and February showed a gradient between low concentrations at the sediment/lake water interface (August E2: 0.63 mg/L; August E4: 1.80 mg/L; February E2: 0.40 mg/L; February E4: 0.37 mg/L) and a subsurface phosphorous maximum (August E2: 5.6 mg/L at 4 cm depth; August E4: 4.00 mg/L at 6 cm depth; February E2: 6.70 mg/L at 12 cm depth; February E4: 5.30mg/L at 12 cm depth) ( Figure 9). The diffusive phosphorous fluxes from sediment pore water (Equation (1)) ranged in August between 0.8 mg/m 2 /d (E2) and 1.9 mg/m 2 /d (E4) and between 0.2 mg/m 2 /d (E2) and 0.7 mg/m 2 /d (location E4) in February.  A peak in lanthanum pore water concentrations can be observed in February at locations E2 (7.40 µg/L at 3 cm depth) and E4 (30 µg/L at 1 cm depth) ( Figure 11). The pore water profiles in August show only a slight enrichment of lanthanum at E2 (1 µg/L at 1 cm and 2 cm depths) relative to the background concentrations (range~0.01-0.5 µg/L) derived from deeper sediments. No pore water lanthanum peak was observed at E4 in August.

Phosphorous in Groundwater
The chemical composition of all groundwater samples taken from the wells is presented as supplementary material in Table S1. Phosphorous concentrations ranged between 0.42 mg/L (GW5028) and 0.70 mg/L in GW979 ( Figure 12). Samples obtained from the banks of the lake by means of push-point sampling showed phosphorous concentrations between 0.34 mg/L and 0.70 mg/L. The Dove-Elbe river water showed phosphorous concentrations of 0.49 mg/L in August 2018 and 0.03 mg/L in February 2019 (note that no phosphorous data are available for April 2018) ( Figure 12).

Water Stable Isotopes
The isotopic signatures of water samples from the Dove-Elbe, from the wells GW11018 and GW5028, and the three push-point samples (P4, P5, P6) plot in the δ 18 O/δ 2 H diagram close to both the Global Meteoric Water Line (GMWL, [39]) and Local Meteoric Water Line (LMWL, data from Global Network of Isotopes in Precipitation, location Cuxhaven) ( Figure 13). These signatures differ from the isotopic signature of the lake water indicating that the lake is neither directly hydraulically connected to the Dove-Elbe nor to the groundwater represented by wells GW11028 and GW5028. Instead, the plot of the isotopic signature of the lake water is close to that of the groundwater sampled at well GW979. This suggests a water exchange between the aquifer and the lake in this area, i.e., at the northern bank of the lake. The quantification of this exchange will be discussed in the following based on a D/O mass balance (see Section 2.5, Equation (2)). The parameters of this mass balance are summarized in Table 2.   (2) we used meteorological data (precipitation, P (mm); air temperature, T ( • C); and relative humidity, h (%); obtained from the German Weather Service weather station Fühlsbüttel located 18 km from Lake Eichbaumsee), and the D/O signatures of the local precipitation (obtained from the GNIP (Global Network of Isotopes in Precipitation), www.iaea.org/services/networks/gnip (accessed on 16 December 2020); Cuxhaven, located about 100 km from Lake Eichbaumsee). Monthly D/O means were available for the years 1978 through 2012. Precipitation and evaporation (mm/d) were adapted to the surface area of the lake. The determination of the average isotopic signature of lake water (δ L ) was based on the depth weighted average isotopic water signatures determined in April (average of profiles E2 and E4:  (Equation (4)) was determined following Petermann et al. [16] by varying the seasonality factor k to fit the local evaporation line (LEL) (Figure 14). The LEL was created based on lake water samples and the groundwater sample taken from well GW5028. The isotopic signature of δ E ranges between −17.3 ‰ and −26.9 ‰ for δ 18 O and between −94.5 ‰ and −116‰ for δ 2 H (Table 2).  Table 2).
Based on Equation (2) the groundwater discharge (G i ) ranged between 1900 and 4100 m 3 d in April and between 2100 and 4800 m 3 /d in August (Table 2). Lake water exfiltration into the aquifer (G o ) ranged between 1660 and 3800 m 3 /d in April 2018 and between 1600 and 4200 m 3 /d in August (Table 2).
From the lake water exfiltration rate (G o ) and the lake water volume a mean water residence time (τ = V/G o; [16]) of between 1.0 and 2.5 years was calculated for the lake.
It should be noted that the results of the isotopic mass balance presented above are associated with uncertainties and limitations. For instance (and as mentioned earlier), the mass balance model assumes a constant lake volume, which may however change between summer and winter. The isotopic signature of precipitation was derived from Cuxhaven, a location close to the seaside about 100 km away from the study area. Hence, this isotopic signature may not be fully representative of the situation at the lake. The same situation holds for humidity, temperature, and precipitation data that was obtained from a weather station 18 km distance from the lake. Further uncertainties may derive from seasonal and depth-dependent variable isotopic signatures of the groundwater in the hydraulically connected aquifers.

Changes in Lake Water Phosphorous Concentrations
The physical and chemical parameters of the lake water column show an annual cycle typical of most temperate lakes with a stratification in spring/summer accompanied by anoxic conditions in the bottom layer, i.e., the hypolimnion, which is characterized by an increase in water phosphorous concentrations with water depth (Figures 3-5). In order to investigate the long-term trend of the phosphorous content of the lake since the last application of Bentophos ® in autumn 2013 we used the lake monitoring data determined by local authorities [28,29,40]. This data is based on seasonal water column phosphorous measurements. The lake water phosphorous content was derived by summing the phosphorus content of each water level (i.e., the product of phosphorous concentration in a water depth times the lake water volume at this depth).
Since 2014 there has been a continuous increase of the lake phosphorous content from 226 kg in autumn 2014 to 935 kg in autumn 2019 ( Figure 15) which translates to an average increase of~140 kg/a. This increase has a clear trend with higher increases between the autumns 2014/2015 (309 kg) and 2015/2016 (160 kg) and relative lower ones between the autumns 2016/2017 (68 kg), 2017/2018 (71 kg), and 2018/2019 (61 kg), respectively. The winter lake water circulations which are expected to remove some of the phosphorous from lake waters show no clear trend and remove on average 166 kg/y of phosphorous. The possible phosphorous sources and sinks, as they result from the presented study, are discussed in the following.

Internal Cycling as a Phosphorous Source
Phosphorous removal from the lake water body is believed to be closely related to the precipitation of Fe(III) hydroxides, which scavenge phosphorous from the water column during full lake water oxygenation in winter [41]. In contrast, during anoxic conditions in summer the reverse process of reductive dissolution of Fe(III) hydroxides releases phosphorous from the sediments. Observations from Danish lakes suggest that if surface sediments show a Fe/P ratio of >15 (by weight) the iron pool is high enough to effectively bind phosphorous during aerobic conditions [42]. For instance, in the artificial Lake Hinterm Horn, which is located about 3 km southeast of Lake Eichbaumsee, the relatively high sedimentary Fe/P ratio of 17.8 is believed to be the main reason for year-around oligotrophic lake conditions [29]. In Lake Eichbaumsee, in contrast, we observed in surface sediments lower Fe/P ratios (6.40-10.0) (cf. Table S2). These ratios are in the same range as the ratio determined in surface sediments in 2016 (Fe/P = 8.90) [29]. Our observed water column seasonal cycles of iron and manganese follow those of phosphorous ( Figures 5 and 6), suggesting an important role of iron and manganese in phosphorous recycling processes. However, a recent study challenges the role of manganese in the removal/liberation processes of phosphorous highlighting the controls of iron redox reactions in phosphorous cycling [43]. Hupfer and Lewandowski [44] pointed out that the retention and mobilization of phosphorous in sediments are very complex processes and dependent on several factors such as the availabilities of oxygen, iron oxides, sulfates/sulfides, carbonates, and phosphorous-binding minerals as well as redox-insensitive binding systems (e.g., Al(OH) 3 and unreducible (Fe(III)). In our study no data on the different phosphorous speciations was available, thus these potentially occurring processes could not be investigated in detail. Nevertheless, it seems that Lake Eichbaumsee sediments retained phosphorous effectively, as sediment phosphorous concentrations increase with decreasing sediment depths, i.e., sediments are a net phosphorous sink (Figure 7). If sediments were to be a net phosphorous source, gradually decreasing sediment phosphorous concentrations towards the sediment surface would be expected, i.e., a pattern similar to that of pore water phosphorous (Figures 7 and 9).
In our study, we observed a diffusive phosphorous flux from sediment pore waters to the lake in February 2019 (i.e., during late winter) with phosphorous flux rates between 0.2 mg/m 2 /d (E2) and 0.7 mg/m 2 /d (E4). During this time no oxygen was present in the lowermost water column at location E4 (water depth 16 m) (Figure 4), suggesting that anoxic conditions prevail in the deep water column during winter. At location E2, characterized by a slightly shallower water depth (14 m), oxygen concentrations were constant throughout the water column; however, we still observed a small phosphorous flux from sediments at this location as well. It is possible that diffusive oxygen uptake by sediments, e.g., due to aerobic respiration, may cause steep oxygen gradients close to the sediment/water interface [45], which were not resolved by our water-column sampling. The presented findings indicate that, at least in the deepest parts of the lake, phosphorous is continuously released from the sediments throughout the winter.
Our results of diffusive phosphorous fluxes from sediments are based only on two pore water profiles from two locations representing both summer and winter situations. Thus, it remains unclear how spatially and temporarily representative these results are. Our study did not cover shallower parts of the lake where, e.g., during winter times, presumably phosphorous uptake by sediments will occur (see Section 4.1). Thus, further data is desirable for a more detailed investigation of the seasonal balance of the phosphorous internal cycling in the studied lake. Nevertheless, assuming that the determined phosphorous pore water fluxes are spatially and temporarily representative for the area of the hypolimnion (154,000 m 2 ; [28]), the total phosphorous fluxes to the lake range between 128 g/d and 300 g/d in August (48 kg/a and 110 kg/a) and between 32 g/d and 100 g/d in February (11 kg/a and 37 kg/a). These rough estimates show that even under a (unlikely) continuation of the highest phosphorous flux (110 kg/a) throughout a full seasonal cycle this would be not sufficient to fully explain the observed water column phosphorous increase, for instance the phosphorous increase of 344 kg between April and November 2018 (=218 days, i.e., 576 kg/a) ( Figure 15).
The observed trend of relatively higher lake water phosphorous increase in the seasons 2014 to 2016 compared to the following years (see Section 4.1) may be explained by a reduced phosphorous release from the sediments after 2016. However, the highest lake water phosphorous increase (relative to the previous year) was observed in summer 2015 ( Figure 15), which was after Bentophos ® phosphorous removal in November 2013. Bentophos ® is expected to remove and fix phosphorous efficiently [46] and the observed sedimentary phosphorous concentration peaks (corresponding to peaks in lanthanum) most probably reflect this enhanced phosphorous accumulation resulting from the Bentophos ® applications (Figure 7). It is not likely that this fixed phosphorous was recycled again as the discussed pore water phosphorous concentration profiles do not indicate any additional phosphorous mobilization from these sediment layers characterized by phosphorous and lanthanum concentration peaks (Figures 9 and 11).
In conclusion, the discussions above show that the observed lake water phosphorous increase cannot convincingly be explained phosphorous internal cycling only.

Groundwater-Borne Phosphorous Flux
As mentioned above, the collected data suggests that according to the water stable isotope mass balance there is an exchange between the aquifer represented by well GW979 and the lake water body. This exchange causes phosphorous flux from the lake to the aquifer and vice versa. Using the range of groundwater exfiltration into the aquifer (G o ) ( Table 2) and the average phosphorous concentration of the lake water (0.49 mg/L; average of November and February) the phosphorous flux to the aquifer ranged between 0.8-1.1 kg/d.
In contrast, combining the range of the groundwater discharge to the lake (G i ) with the phosphorous concentration of the aquifer groundwater at well GW979 (0.7 mg/L), results in phosphorous fluxes to the lake via groundwater discharge of between 1.3-3.3 kg/d. Thus, the net phosphorous infiltration to the lake amounts to 0.4-1.1 kg/d or 150-390 kg/a. It should be noted that the phosphorus flux calculated based on the water stable isotope balance is a long-term average flux (note, the lake water residence time is~1 to 2.5 years) and thus does not reflect short-term seasonal changes in the groundwater phosphorous supply. Nevertheless, the isotope mass balance suggests that groundwater discharge to the lake is in the same range or higher than the sediment-derived phosphorous flux.
The observed trend of a lower total phosphorous increase in the lake water in autumn since 2016 (see Section 4.1) can be explained by the difference between lake water and groundwater phosphorous concentrations. The difference of the lake water to the groundwater concentration of 0.70 mg/L was highest in springs 2014 and 2015 (lake phosphorous 0.15 mg/L) and was less in spring 2019 (~0.43 mg/L). With decreasing differences in the phosphorous concentrations between the two water bodies one expects the phosphorous increase in Lake Eichbaumsee to reach a plateau value. Without any direct observations or direct measurements of phosphorous-rich groundwater discharge to the lake, it is difficult to exactly define the origin of phosphorous in the lake. Groundwater, while seeping through lake sediments, may mobilize phosphorous from iron oxides, phosphorous-bearing minerals, and organic matter buried in lake sediments [10,13]. Although phosphorous was long considered to be immobile in groundwater, there is increasing evidence that the eutrophication of lakes is driven by groundwater discharge rich in phosphorous and that this may be an important albeit so far neglected source [47][48][49][50][51]. For instance, studies of the eutrophic Lake Arendsee (Germany) revealed that groundwater discharge was responsible for more than 50% of the total phosphorus supply to the lake [48]. The importance of groundwater-borne phosphorous may also hold for oligotroph lakes [49].
Phosphorous in groundwater may derive from natural and anthropogenic sources [52,53]. In the presented study, phosphorous groundwater concentrations are below the EU drinking water standard (2.20 mg P L −1 ), thus there is no indication of any substantial impact of anthropogenic phosphorous sources. Identifying such sources is challenging, and analyses of the phosphorous oxygen isotope composition may be a powerful tool to be used in further studies on the lake phosphorus cycle [54].

Conclusions
More than 70% of the lakes in Germany (n = 732) do not achieve the "good ecological status" required by the European Water Framework Directive [55,56], with excess nutrient loads being the major cause for the deterioration of the quality of lake waters. Thus, there is a general need for restoration measures which are concentrated on the internal phosphorus cycle only. The presented study was carried out at the eutrophic Lake Eichbaumsee, which is characterized by poor water quality caused by high phosphorous concentrations. Despite several remediation attempts, e.g., water aeration or removal of dissolved phosphorous by co-precipitation with aluminum sulfate and lanthanum-modified bentonite (Bentophos ® ), lake water phosphorous concentrations have increased significantly over the years. Assuming that our determined sediment pore water fluxes (11-110 kg/a) are temporally and spatially representative, the lake internal phosphorous source (i.e., phosphorous released from sediments) is not sufficient to explain the observed long-term phosphorous increase in the water column. Furthermore, the results do not indicate any remobilization of sedimentary phosphorous co-precipitated with Bentophos ® . Based on a D/O mass balance we show that groundwater-borne phosphorous flux is a significant but so far not considered phosphorous source for the water body of Lake Eichbaumsee. This groundwater source needs to be considered in any successful remediation attempts of the lake, a situation which is most probably also relevant for other eutrophic groundwater-fed lakes.