Fate of organic micropollutants in the hyporheic zone of a eutrophic lowland stream: Results of a preliminary field study

Science of the Total Environment 409 (2011) 1824–1835

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Science of the Total Environment j o u r n a l h o m e p a g e : w w w. e l s ev i e r. c o m / l o c a t e / s c i t o t e n v

Fate of organic micropollutants in the hyporheic zone of a eutrophic lowland stream: Results of a preliminary field study Jörg Lewandowski a,⁎, Anke Putschew b, David Schwesig c, Christiane Neumann d, Michael Radke d a

Leibniz-Institute of Freshwater Ecology and Inland Fisheries, Department Ecohydrology, Müggelseedamm 310, 12587 Berlin, Germany Technical University Berlin, Department of Water Quality Control, Strasse des 17. Juni 135, 10623 Berlin, Germany IWW Water Centre, Moritzstr. 26, 45476 Mülheim an der Ruhr, Germany d University of Bayreuth, Department of Hydrology, BayCEER, Universitätsstr. 30, 95440 Bayreuth, Germany b c

a r t i c l e

i n f o

Article history: Received 28 May 2010 Received in revised form 16 January 2011 Accepted 18 January 2011 Available online 23 February 2011 Keywords: Pharmaceuticals Ibuprofen Hyporheic zone Self-purification capacity Gadolinium Borate Hydrology Biogeochemistry

a b s t r a c t Many rivers and streams worldwide are impacted by pharmaceuticals originating from sewage. The hyporheic zone underlying streams is often regarded as reactive bioreactor with the potential for eliminating such sewage-born micropollutants. The present study aims at checking the elimination potential and analyzing the coupling of hydrodynamics, biogeochemistry and micropollutant processing. To this end, two sites at the lowland stream Erpe, which receives a high sewage burden, were equipped and sampled with nested piezometers. From temperature depth profiles we determined that at one of the sites infiltration of surface water into the aquifer occurs while exfiltration dominates at the other site. Biogeochemical data reveal intense mineralization processes and strictly anoxic conditions in the streambed sediments at both sites. Concentrations of the pharmaceuticals indomethacin, diclofenac, ibuprofen, bezafibrate, ketoprofen, naproxen and clofibric acid were high in the surface water and also in the subsurface at the infiltrating site. The evaluation of the depth profiles indicates some attenuation but due to varying surface water composition the evaluation of subsurface processes is quite complex. Borate and non-geogenic gadolinium were measured as conservative wastewater indicators. To eliminate the influence of fluctuating sewage proportions in the surface water, micropollutant concentrations are related to these indicators. The indicators can cope with different dilutions of the sewage but not with temporally varying sewage composition. © 2011 Elsevier B.V. All rights reserved.

1. Introduction The hyporheic zone—the transition zone between surface water in streams and groundwater (Runkel et al., 2003)—is a key compartment of the hydrosphere. It is of utmost importance for maintaining the ecological function of running waters and a natural reactor taking main responsibility for the impressive self-purification capacity of lotic systems, from small brooks to large rivers. The hyporheic zone is also a barrier against contamination of near-surface aquifers, which are essential for the production of drinking water. Its ecological service is provided and sustained by the interaction of physical (e.g., transport of water and solutes), chemical (e.g., chemical reactions, sorption), and biotic processes (e.g., microbial transformation, bioturbation) by diverse and active hyporheic communities (Krause et al., 2009).

⁎ Corresponding author. Tel.: +49 30 64181 668; fax: +49 30 64181 663. E-mail addresses: [email protected] (J. Lewandowski), [email protected] (A. Putschew), [email protected] (D. Schwesig), [email protected] (C. Neumann), [email protected] (M. Radke). 0048-9697/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.scitotenv.2011.01.028

Numerous investigations on nitrogen, phosphorus and organic carbon processing in rivers highlight the high intrinsic potential of hyporheic zones as efficient bioreactors. As in other environmental systems, the cycling of nutrients and other major water compounds is closely coupled. For example, phosphorus cycling is impacted by the cycling of iron, aluminium, calcium, and sulphur (Hendricks and White, 2000; Reddy et al., 1999; Roden and Edmonds, 1997; House, 2003). While the reducing milieu in the hyporheic zone favours the unwanted mobilization of phosphate, it also favours the elimination of nitrate (Lewandowski and Nützmann, 2010). Denitrification may occur even in well-oxygenated environments due to the presence of anaerobic microhabitats (Birgand et al., 2007; Hendricks and White, 2000; Mulholland and DeAngelis, 2000). Over a range of diverse headwater streams, for example, 70–80% of ammonium removal from the flowing water was attributed to uptake and subsequent consumption in the streambed sediments (Peterson et al., 2001). Beside nutrients, organic micropollutants such as pharmaceuticals and personal care products are present in streams and rivers. The concern about their presence is mainly related to potential adverse effects on environmental systems, their bioaccumulation potential, and to human toxicology when surface water is used for the production of drinking water. In spite of large efforts to minimize

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the emission of such compounds into the environment, they are ubiquitous in surface waters. Their major sources to streams are emissions from wastewater treatment plants (WWTPs). Other sources contributing to the environmental burden are for example sewer overflows, and diffuse sources like runoff from agricultural or industrial areas. In contrast to non-polar organic contaminants like polycyclic aromatic hydrocarbons (PAH) or polychlorinated biphenyls (PCB), sorption of these emerging micropollutants to sewage sludge is less efficient due to their higher polarity (Kalsch, 1999). Biodegradation of many micropollutants in WWTPs is comparatively inefficient (Zwiener and Frimmel, 2003). The assessment of pharmaceuticals by standardized tests indicated low biodegradation (Al-Ahmad et al., 1999), and thus these emerging micropollutants are frequently referred to as pseudo-persistent in the environment (Daughton, 2003). However, studies performed under experimental conditions more resembling streams or studies on bank filtration and artificial groundwater recharge show that the same compounds considered persistent in WWTPs are transformed in river sediments or other porous matrices (Schittko et al., 2004; Löffler et al., 2005; Gruenheid et al., 2008; Kunkel and Radke, 2008; Schulz et al., 2008; Radke et al., 2009). This apparent contradiction is mainly due to two reasons: i) the residence time in WWTPs is too short to allow efficient biodegradation while in the porous space of river sediments or aquifers the residence time can be much longer, and ii) the microbial community in environmental systems is much more diverse than in WWTP. In river water, transformation processes in the absence of sediment are comparatively inefficient for many micropollutants (Kalsch, 1999; Radke et al., 2009). Thus, the hyporheic zone is hypothesized to be a key compartment with major influence on the environmental fate of organic micropollutants in lotic systems. To study the fate of sewage-borne organic micropollutants and nutrients in the stream and the hyporheic zone, conservative sewage indicators are helpful. Borate was frequently used as such an indicator due to its predominantly anthropogenic origin in many streams. However, the amount of perborate in detergents is decreasing for more than a decade now and thus borate is nowadays less suitable (Neal et al., 2010). In recent years, the use of the rare earth element gadolinium (Gd) as a sewage indicator in hydrology has been discussed (Verplanck et al., 2005). Gd compounds have been used as contrasting agent in clinical diagnosis (magnetic resonance imaging) since 1988 (Kümmerer and Helmers, 2000) and are currently used in about 150 million annual applications worldwide. Gd is usually administered as a complex (up to 0.3 mmol kg−1 body weight) and excreted unmetabolised within a few hours (Kümmerer and Helmers, 2000). Such organic Gd complexes are able to reach the surface water systems because they are stable enough to pass nearly unaffected through common WWTPs (Möller et al., 2003; Künnemeyer et al., 2009). During the last decade, a number of studies from various countries have reported anthropogenic Gd in rivers (Bau and Dulski, 1996; Elbaz-Poulichet et al., 2002; Möller et al, 2003; Morteani et al., 2006). Due to its stability, low sorption tendency and low geogenic background, the use of anthropogenic Gd as a conservative indicator of urban wastewater in rivers has been proposed (Verplanck et al., 2005). Furthermore, even in the case of geogenic Gd present, the anthropogenic fraction of Gd can be calculated from the concentration pattern of the other rare earth elements (REE) in a water sample. Most previous studies on hydrology and biogeochemistry of the hyporheic zone were conducted in small headwater streams. Water quality in headwaters is usually more pristine than in lowland streams. Due to lower flow velocities and more eutrophic conditions, bed sediments of lowland streams are usually much finer with higher organic matter content than those occurring in headwaters. Often the redox milieu in streambed sediments of lowland streams is anoxic while headwater streambed sediments are usually aerobic. Due to the texture of the streambed sediment in headwaters there is an intense exchange between surface water and the hyporheic zone. Flow

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velocities in the hyporheic zone are high compared to lowland streams. It is widely unknown whether unfavourable hydraulic conditions spoil the importance of the hyporheic zone in lowland streams. On the one hand, lowered exchange rates and lower flow velocities in the hyporheic zone decrease the percentage of water passing through the hyporheic zone. On the other hand, this increases the reaction time in the hyporheic zone. The underlying question is whether transformation of micropollutants in the hyporheic zone is limited by transport (of micropollutants and/or reaction partners) or kinetics. Gücker and Pusch (2006) investigated whether the paradigm of effective nutrient retention, also known as self-purification concept, derived from pristine headwater streams (e. g. Peterson et al., 2001), holds true for lowland streams with excessive nutrient loads like the stream Erpe. Haggard et al. (2001) and Marti et al. (2004) reported low load-specific nutrient retention efficiencies in such streams with high sewage burden and the study of Gücker and Pusch (2006) at the lowland stream Erpe confirms this conclusion. However, an overload of the system hypothesized as cause for the low relative efficiency of nutrient retention cannot be assumed for micropollutants. Thus, we want to check whether the hyporheic zone is a compartment with paramount responsibility for transformation of micropollutants in lowland rivers. The aims of the present preliminary study are (1) to analyze the coupling of hydrodynamics, biogeochemistry and micropollutant processing in the hyporheic zone of a lowland river, (2) to evaluate whether borate and anthropogenic gadolinium are suitable indicators for the proportion of wastewater in the hyporheic zone, and (3) to investigate the hypothesis that the hyporheic zone is an important compartment for transformation of micropollutants in the stream Erpe. Previous research of nutrient retention in the hyporheic zone (e.g. Gücker and Pusch, 2006) was usually based on tracer tests conducted in the overlying water. In tracer tests surface storage and hyporheic exchange are lumped together in a single storage zone, so it is often difficult to distinguish between surface and subsurface storage (Cardenas, 2006; Runkel et al., 2003; Wörman et al., 2007). Furthermore, hyporheic processes and their interaction cannot be studied in detail on the scale of tracer tests, and downscaling from larger scales is impossible without prior knowledge on the individual processes, especially due to the large heterogeneity of the hyporheic zone (McClain et al., 2003). Therefore, in the present study we investigate the hyporheic processes on a decimeter scale directly in the hyporheic zone. Knowledge on the aforementioned aims are of major importance since many streams and rivers in densely populated areas are sewage-polluted lowland rivers while most studies on hyporheic zone processes were conducted in pristine headwater streams so far. We have chosen the stream Erpe for our investigations because it represents a typical lowland stream with high sewage burden, and previous investigations of nutrient retention by Gücker and Pusch (2006) and Gücker et al. (2006) provide a useful basis for our study. 2. Materials and methods 2.1. The study site The lowland stream Erpe, also known as Neuenhagener Mühlenfließ, is located at the eastern edge of Berlin. It is polluted by intense agriculture in its catchment as well as several point sources such as septic tank spillways, small private and municipal WWTPs and the large WWTP Münchehofe (Köhler et al., 2002). The latter WWTP has a dry weather capacity of 42,500 m3 d−1, which equals to 220,000 population equivalents (PE). The treatment technology of the WWTP includes denitrification and chemical P-precipitation steps. Two sites at the stream Erpe were studied. Site ABC (52°29′00.24″N, 13°38′ 28.27″E) is located 150 m downstream of the confluence with the WWTP Münchehofe effluent; site DF (52°28′19.18″N, 13°37′22.61″E)

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is located 2 km downstream of the first site. The floodplain of the Erpe at both studied sites and in between is several 100 m wide. The Erpe was straightened and incised in the 1970s (Gücker and Pusch, 2006). Submerged macrophytes (Potamogeton pectinatus L. and Sparganium emersum Rehmann) growing from May to September dominate primary production in the Erpe and control sediment storage in the stream (Heppell et al., 2009). The local water management authorities remove these macrophytes annually in autumn, thereby restoring the initial streambed morphology (Gücker et al., 2006). The soft-bottom lowland stream Erpe has predominantly clogged and anaerobic fine-sandy siliceous sediments with high organic carbon and nutrient contents (Gücker et al., 2006). To investigate stream hydrodynamics as potential controls of nutrient uptake, shortterm conservative tracer addition experiments were conducted by Gücker and Pusch (2006) and evaluated with OTIS-P, a onedimensional advection-dispersion model that includes transient storage and lateral inflow (Runkel, 1998). Absolute nutrient uptake rates (ammonium 0.076 to 0.517 gN m−2 d−1, nitrate 5.0 to 12.9 g N m− 2 d− 1, phosphate 0.18 to 0.75 g P m−2 d−1; Gücker and Pusch, 2006) were high while relative uptake rates were low, indicating that nutrient uptake in the hyporheic zone only compensated a minor proportion of the high loads in the stream, resulting in long nutrient uptake lengths. Present-day loadings of efficiently treated wastewater do not cause extended stream reaches to become dominated by massive heterotrophic processes. This is different as compared to crudely treated wastewater with high loadings of organic matter, which formerly prevailed in sewage-impacted water bodies (Hynes, 1974) and favoured heterotrophic microbial metabolism. Instead, under current conditions both autotrophic and heterotrophic processes are stimulated (Gücker et al., 2006; Gücker and Pusch, 2006): gross primary production of 27.5 g dissolved oxygen m−2 d−1 and community respiration of 59.2 g dissolved oxygen m−2 d−1.

2.2. Sampling technique Nested streambed piezometers similar to those described by Rivett et al. (2008; Fig. 1(e) therein without washers (N) and nut (O)) were used to sample hyporheic water or groundwater, respectively. Each piezometer consisted of a color-coded polytetrafluorethylene (PTFE) tube (inner diameter 2 mm, outer diameter 4 mm). At the lower end a filter screen was made from a 10 cm long PTFE tube (inner diameter 4 mm, outer diameter 6 mm). Approx. 40 small holes were cut into that tube to increase the filter area. The filter tube was wrapped by gauze (mesh size 0.2 mm), which was stitched with a fibre and fixed with two cable clips. Five piezometers were attached with cable clips to

Normalised REE concentration (10-6)

Gd*measured Gd*expected

8

6 Gd*measured 4 Interpolated function

Gd*expected

2.3. Temperature-based measurements of surface-subsurface water exchange Sediment temperatures at both sites were measured with a stainless steel temperature probe containing 8 temperature sensors (5, 10, 15, 25, 45, 70 and 95 cm beneath the sediment–water interface; 5 cm above the interface; type K thermocouples), which are separated by a plastic inner core. The probe was driven 1 m into the sediment and after an equilibration phase of 10–20 min the temperatures were recorded with a thermologger (Voltcraft, K204, resolution: 0.1 K, Accuracy: 0.2 K). Exchange across the sediment–water interface was then calculated using an analytical solution of the one-dimensional heat and fluid flow transport equation for steady state conditions (Bredehoeft and Papadopolus, 1965): ! qz ρw cw exp z −1 kfs T ðzÞ−T0 ! = TL −T0 qρ c exp z w w L −1 kfs

ð1Þ

with: TL—constant temperature at the lower boundary [°C], T0— temperature at the sediment–water-interface [°C], T(z) temperature at depth z [°C], qz—vertical Darcy velocity [L m−2 d−1], ρwcw—heat capacity of water (4.2.106 J m−3 K−1), kfs—thermal conductivity of the saturated sediment (2 J s−1 m−1 K−1, value from Stonestrom and Constantz, 2003), z—depth of sensor [m], L—depth of constant temperature [m]. Eq. (1) is solved for T(z). To obtain an optimum fit of the simulated to the measured temperature profile, the value of qz was adjusted by minimizing the mean of the squared differences between the measured and the modelled temperatures. Due to the lack of knowledge of the constant groundwater temperatures, Eq. (1) was solved with the measured temperature at a depth of 95 cm.

10 Gdanomaly =

an aluminium rod (length: 1 m, diameter: 5 mm) with filter screens in 0–10, 10–20, 25–35, 40–50 and 90–100 cm. For installation of the nested piezometers a metal tube (inner diameter 19 mm) was pushed or hammered into the sediment. A nested piezometer was inserted into the metal tube and the metal tube was pulled out of the sediment. An aluminium tip loosely attached to the metal tube was lost in the sediment when pulling out the metal tube. Three piezometer nests (ABC and DEF) were installed at the sampling sites. Nested piezometers B and E were installed in the centre of the stream, A and D 1.5 m apart from the right bank (in downstream direction) and C and F 1.5 m apart from the left bank. Piezometer E was lost prior to sampling. Two to four days after installation hyporheic water was sampled with a multi channel peristaltic pump (Ismatec IPC 24) on July 8, 2009. The tubes of the nested piezometers were extended with PTFE tubes to reach the bank. Sampling was conducted simultaneously for all five ports of each nested piezometers and all nested piezometers at a location (ABC or DF). After discarding twice the dead volume of the tubes water samples were collected at a low flow rate of approximately 250 mL h−1. In total 900 mL sample volume were collected and immediately split into the different sub-samples required for analysis. At two ports flow rates were too low to collect any sample (C 0–10 cm, C 10–20 cm) and at 6 ports low flow rates reduced the number of analyzable parameters (A 10–20 cm, D 0–10 cm, D 10–20 cm, D 40–50 cm, F 25–35 cm, and F 40–50 cm).

2

2.4. Major ions 0 57

59

61

63

65

67

69

71

73

REE atomic number Fig. 1. Scheme how to calculate a Gd anomaly from shale-normalised REE data.

Samples were filtered immediately after sampling (0.45 μm, cellulose acetate) and cooled until analysis. To avoid a loss of ammonium, dissolved iron and dissolved manganese the corresponding samples were acidified with 2 M HCl to a pH of

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Table 1 CAS-No., pka and structure of the acidic pharmaceuticals (Huber et al., 2003; Scheytt et al., 2005; Harding et al., 2009). Compound

CAS No.

pKa (Reference)

Bezafibrate

41859-67-0

3.6 (Huber et al., 2003)

Clofibric acid

882-09-7

3.20 (Scheytt et al., 2005)

Diclofenac

15307-86-5

4.15 (Harding et al., 2009)

Ibuprofen

15687-27-1

4.45 (Harding et al., 2009)

Indomethacin

53-86-1

4.50 (Harding et al., 2009)

Ketoprofen

22071-15-4

4.45 (Harding et al., 2009)

Naproxen

22204-53-1

4.15 (Harding et al., 2009)

approximately 2. Nitrate and sulphate were measured by ion chromatography (Dionex, DIN EN ISO 10304-1). Ammonium was determined photometrically by a segmented flow analyzer (Skalar, EN ISO 11732). Samples for metal analysis were measured by atomic absorption spectrometry (Perkin Elmer, DIN 38406). Total dissolved phosphorus (TDP) was determined photometrically by the molybdenum blue method after digestion. Boron was quantified according to DIN 38405-17. The analysis is based on the reaction of boron with azomethin-H in a buffered solution (pH 7.3). The reaction product of azomethin and boron was measured by a spectrophotometer (Perkin Elmer, Lamba 12) at a wavelength of 414 nm.

Structure

2.5. Rare earth elements (REE) analysis and calculation of non-geogenic Gd Samples for REE analysis were acidified by adding 1 mL nitric acid (65%, suprapure, Merck) per 50 mL sample volume after filtration (0.45 μm). Analysis of REE was performed by ICP-MS (Elan DRC II, PerkinElmer Sciex) in the standard mode. All REE from atomic order numbers 58 (Ce) to 71 (Lu) were measured in one analytical run. Calibration solutions were prepared from a REE multielement standard (stock solution 1000 mg L−1 from HighPurity Standards, Germany). A ten-point linear calibration function (calibration range 10–500 ng L−1)

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2.6. Organic micropollutants

Table 2 Parent ions, recorded product ions and collision energy. Compound

[M–H]−

Product ion I

Product ion II

Collision energy

Naproxen Ketoprofen Bezafibrate Ibuprofen Diclofenac Indomethacin Clofibric acid

229 253 360 205 294/296 356/358 213/215

185 209 274 161 250 312 127

170

7/15 8 16/30 8 13/13 8/8 13/13

154 252 314 129

The pharmaceuticals indomethacin, diclofenac, ibuprofen, bezafibrate, ketoprofen, naproxen and clofibric acid were analyzed by LCMS/MS after solid phase extraction. The molecular structures, CAS registry numbers, and pKa values of these compounds are summarized in Table 1. The samples were filtered (0.45 μm, cellulose acetate) and then enriched with an AutoTrace SPE Workstation (Zymarck, Idstein, Germany) and disposable Oasis HLB (60 mg, Waters) cartridges. Before use the SPE cartridges were cleaned and activated with 7 mL methanol/acetonitrile (60:40, v:v) and 7 mL ultra pure water; subsequently 100 mL sample were extracted. After extraction the cartridges were dried with nitrogen. The compounds adsorbed on the SPE material were eluted with 7.5 mL methanol/acetonitrile (60:40, v:v). The extracts were concentrated to a final volume of 1 mL using a TurboVap II (Zymarck, Idstein, Germany). Quantification was based on a calibration in matrix (water collected from the Erpe directly upstream of the WWTP Münchehofe). Standards were added in 6 different concentrations (0.02 μg L−1–1 μg L−1) and subsequently extracted and treated as described above. The extract from the WWTP was diluted 1:20 before analysis. LC separation was carried out using a HP System (Hewlett-Packard series 1100, Waldbronn, Germany) comprising a vacuum solvent degassing unit, a binary pump and an autosampler. Separation was done on a Varian Monochrom MS column (5 μm, 100 × 2 mm). The solvents were A: ultra pure water: methanol (80:20, v:v) and B: ultra pure water:methanol (5:95, v:v). The gradient elution was as follows: 3 min 50% B, from 50% B to 85% B in 9 min which was held for 4 min, to 50% B in 1 min which was held for 7 min. The injection volume was 20 μL. The LC was coupled to a triple-quadrupole mass spectrometer (Quattro-LC; Micromass, Manchester, UK). Nitrogen generated from pressurized air (Whatmann, Haverhill, USA) was used as drying gas (900 L h-1) and nebulising gas (85 L h−1). The desolvation temperature was 180 °C and the source temperature 120 °C. Negative-ion electrospray ionization was used. The capillary voltage was set to 3 kV. The compounds were detected in the multi reaction monitoring mode (MRM). Argon 5.0 was used as collision gas. The pressure in the collision cell was 1.3 × 10−3 mbar. The parent ions as well as the recorded product ions are shown in Table 2. To facilitate the interpretation of the depth profiles of the micropollutants, we averaged (arithmetic mean) the values of the same depth for all profiles of site 1 (A, B, C). The next step was to normalize the concentrations of each micropollutant to the concentration of that micropollutant in the surface water. Finally, the normalized

was applied for all measured REE, and an independent reference material with a Gd concentration of 50 ng L−1 (Certipur, Merck, Germany) was repeatedly used for quality control after each analytical sequence of 10 samples. Daily quality control routine included also the measurement of REE (500 ng L−1 each) in an artificial surface water (SPS-SW 1, Spectropure, Norway). All standards and reference materials were traceable to NIST standard reference material No. 3100 series. Performance data of the method such as limit of detection and limit of quantification were calculated from the ten-point calibration function according to the German standard DIN 32645 (2008). The procedure to quantify non-geogenic Gd is the calculation of a so-called Gd anomaly. This is based on the (linear or polynomic) interpolation of the expected Gd concentration from the neighbouring REE after shale-normalisation. The calculated Gd anomaly is the ratio of the measured Gd (after normalisation) to the expected (geogenic) Gd (after normalisation). The approach has been described in the geochemical literature (e.g. by Bau and Dulski, 1996; Möller et al., 2003). It is based on the observation that the REE distribution of geogenic materials and of water (that is not anthropogenically influenced) follows a consistent pattern if shale-normalised concentrations are plotted against the atomic number of the elements. By using either linear or polynomial interpolation, the expected value of the (geogenic) normalised Gd can be calculated from the normalised results of the neighbouring REE, and the ratio of measured to expected normalised Gd can be calculated. The general methodology is visualised in Fig. 1. Post-Archaean Australian shale (PAAS) was used for normalisation (McLennan, 1989). The fraction of non-geogenic Gd is calculated from the Gd anomaly and the measured Gd concentration as given in Eq. (2) (In samples with Gd concentrations below the limit of quantification (LOQ) Gdnon-geogenic was also set “bLOQ”):

Gdnon−geogenic ½ng = L =

  Gdanomaly −1 Gdanomaly

Gdmeasured ½ng = L

ð2Þ

River Water

20

Streambed sediment

Sediment depth (cm)

0

-20

-40

-60

-80

-100

15

16

17

18

19

20

21

Site 1A 1C 2E

15

16

17

18

19

20

21

Temperature (°C) Fig. 2. Measured and calculated temperatures at site 1 (left) 2 m apart from the right bank (black circles) and 2 m apart from the left bank (grey squares), respectively, and at site 2 (right). Symbols indicate measured, lines calculated temperatures. Arrow indicates flow direction.

J. Lewandowski et al. / Science of the Total Environment 409 (2011) 1824–1835

River Water

20

Streambed sediment

Sediment depth (cm)

0

-20

-40

-60

-80

-100

1829

0.2

0

0.4

Piezometer A B C D F

100

0

Borate (mg/L)

200

300

400

500

Non-geogenic Gd (ng/L)

Fig. 3. Depth profiles of borate and gadolinium (Gd).

micropollutant concentration was related to the corresponding normalized borate concentration. With the assumption that the concentration of borate in the WWTP effluent is correlated to the concentrations of the micropollutants, this approach allows to exclude variations due to varying concentrations in the surface water or due to varying exchange rates between river and hyporheic zone. Attenuation of micropollutants would be indicated by a ratio b1 while a ratio of 1 reflects no attenuation. Borate concentrations used in these calculations

River Water

20

Streambed sediment

0

-20

-40

Sediment depth (cm)

-60

-80

-100

were corrected for borate background concentrations in the anthropogenically-impacted aquifer of the site (0.1 mg L−1) since the present study focuses on the WWTP effluent in the stream Erpe and not on other anthropogenic borate sources. Alternatively, we normalized the micropollutants to the concentrations in the uppermost sediment layer (port 0–10 cm of the nested piezometers), and we also related the depth profiles of the micropollutants to the Gd profile instead of the borate profile.

0

2

4

6

8

10

0

2

4

6

8

10

0

10

NH4-N(mg/L)

NO3-N(mg/L)

30

20

40

Fe (mg/L)

20

0

-20

-40

Piezometer A B C D F

-60

-80

-100

0

1000

2000

3000

4000

DTP (µg/L)

5000

0

50

100

Sulfat (mg/L)

150

0

1

2

3

4

5

Mn (mg/L)

Fig. 4. Depth profiles of nitrate, ammonium, sulfate, dissolved iron, dissolved manganese and dissolved total phosphorus at both sites including the corresponding surface water concentrations. Some values are missing due to blocked sampling ports.

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River Water

20

Streambed sediment

0

-20

-40

Sediment depth (cm)

-60

-80

-100

0

5

10

15

20

0

100

Clofibric acid (ng/L)

200

0

300

50

Naproxen (ng/L)

100

150

200

Ketoprofen (ng/L)

20

0

-20

-40

-60

-80

-100

0

200

400

600

800

0

Bezafibrate (ng/L)

50

100

150

200

Ibuprofen (ng/L)

250

0

500

1000

1500

2000

2500

Diclofenac (ng/L)

20

0

Piezometer A B C D F

-20

-40

-60

-80

-100

0

50

100

150

200

250

300

350

Indomethacin (ng/L) Fig. 5. Depth profiles of clofibric acid, naproxen, ketoprofen, bezafibrate, ibuprofen, diclofenac and indomethacin at both sites including the corresponding surface water concentrations. Some values are missing due to blocked sampling ports.

3. Results 3.1. Hydrology At site 1, the temperature profiles (Fig. 2) indicated infiltration of stream water into the hyporheic zone. The profiles 2 m apart from the right and the left bank were similar with calculated downward seepage rates (loss of river water) qz of 44.5 L m−2 d−1 and 32.4 L m−2 d−1, respectively. The root mean square error between modelled and calculated temperatures was 0.01 °C. The linear temperature profile at site 2 indicated very little exchange of

surface and groundwater through the hyporheic zone with a calculated upward seepage rate of 6.4 L m−2 d−1. The concentrations of the conservative sewage indicator borate (Fig. 3) were similar in surface and subsurface water at site 1 (A, B, C). A few data points are missing due to blocked sampling ports but in principle all three profiles were quite similar with a concentration of 0.5 mg L−1 in the surface water, an immediate decrease to 0.25 mg L−1 in 0–10 cm depth and a gradual increase to 0.45 mg L−1 in 90–100 cm depth. At site 2 (D, F) borate concentrations in surface water were similar to site 1. In the subsurface the borate concentration decreased rapidly to approximately 0.1 mg L−1.

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20 0 -20 -40

Sediment depth (cm)

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A 0

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4

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8

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5

6

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Fig. 6. Depth profiles of clofibric acid, naproxen, ketoprofen, bezafibrate, ibuprofen, diclofenac and indomethacin (arithmetic mean) at site 1 normalized to the concentrations of each micropollutant in the surface water or the uppermost pore water concentration (0–10 cm). Finally, the normalized micropollutant concentration was related to the corresponding normalized borate concentration (without borate background) or normalized non-geogenic Gd. Vertical grey lines indicate no attenuation.

At both sites the concentration of non-geogenic Gd in the pore water was lower than in the surface water and decreased with streambed depth (Fig. 3). At site 2 (D, F) this decrease was more pronounced than at site 1 (A, B, C). Similar to borate, the decrease of non-geogenic Gd at site D was a little less steep than at site F. The profile C of non-geogenic Gd differed from the other profiles at site 1 since concentrations of non-geogenic Gd were higher than in both other profiles and no decrease with depth was observed. Borate concentrations were 0.620 mg L−1 in the discharge of the WWTP, 0.094 mg L−1 in the stream Erpe upstream of the WWTP's outlet and 0.521 mg L−1 at site 1 150 m downstream of the outlet. Assuming perfect mixing of the surface water passing site 1 we estimated that 81.2% of the surface water originated from the WWTP with endmember-mixing calculation. Surface water concentrations at site 2 were 0.508 mg L−1. Thus, only a slight dilution (2.5%) of the surface water by groundwater exfiltration or surface inflow occurred within the 2 km from site 1 to site 2. In the latter estimation borate concentrations of the additional water were neglected. The analogous calculations for nongeogenic Gd resulted in 78.9% surface water originating from the WWTP and a dilution of 4.2% by additional water within the 2 km from site 1 to site 2. 3.2. Biogeochemistry Fig. 4 shows the concentrations of nitrate, ammonium, sulphate, dissolved iron, dissolved manganese and dissolved total phosphorus in the surface water and along the depth profiles at both sites. Nitrate concentrations in the surface water were high at both sites and below the detection limit throughout the streambed sediment. Sulphate concentrations also decreased at the sediment–water interface but less pronounced than nitrate. Furthermore, profile C at site 1 differed

drastically from the other profiles. Ammonium concentrations were low in the surface water and drastically increased in the uppermost streambed sediment layer; below, there was only little further increase. Both sites were similar regarding the biogeochemical parameters. All dissolved iron and dissolved manganese concentrations were close to the detection limit in the surface water and exhibited a concentration peak in the uppermost streambed layers. Phosphorus concentrations increased at the sediment–water interface and were high throughout the streambed profile, especially at site 1. 3.3. Micropollutants Fig. 5 shows the depth profiles of clofibric acid, naproxen, ketoprofen, bezafibrate, ibuprofen, diclofenac and indomethacin. While surface water concentrations of the micropollutants were similar at both sites (except ketoprofen), subsurface micropollutant concentrations differed between site 1 (A, B, C) and site 2 (D, F) (except indomethacin). Subsurface micropollutant concentrations at site 2 were usually much lower than at site 1. In nearly all depth profiles we observed constant or decreasing micropollutant concentrations with depth. Only the concentration of clofibric acid at site 1 increased from the surface water to the uppermost sediment layer. Fig. 6 shows depth profiles of the micropollutants in relation to the conservative sewage indicator borate or Gd and normalized to the surface water concentrations of the micropollutants or the pore water concentrations in the uppermost sediment layer. Ibuprofen could not be normalized to the surface water concentration since its surface water concentration was below the detection limit. In relation to borate (Fig. 6), naproxen and indomethacin showed a substantial attenuation in the streambed sediment compared to the surface water (Fig. 6C). Borate-normalized ketoprofen and diclofenac were N1 in the

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uppermost sediment layers and decreased with depth. Normalized clofibric acid was clearly N1 in the uppermost sediment layers and decreased also with depth. Depth profiles normalized to the concentrations in the uppermost streambed layer and in relation to borate showed decreasing normalized micropollutant concentrations with depth indicating attenuation (Fig. 6D). Since streambed concentrations of naproxen and indomethacin were quite low (Fig. 5) their normalized profiles (Fig. 6B and D) are quite uncertain. Borate-normalized concentrations of all other micropollutants were higher in the uppermost sediment layer than in the surface water and all of them decreased in the streambed with depth. Micropollutant concentrations normalized to surface water concentrations and in relation to non-geogenic Gd (Fig. 6A) increased in the upper streambed layers and were N1 throughout the depth profile (except naproxen and indomethacin). When normalized to the uppermost streambed layer and relative to non-geogenic Gd (Fig. 6B) some values N1 were observed for all micropollutants; ibuprofen and clofibric acid were the only micropollutants with a normalized concentration ratio b1 in the deepest sampling port. 4. Discussion 4.1. Hydrology Due to fine-sandy organic sediments that prevail in the stream Erpe (Gücker et al., 2006) we would not expect an extended hyporheic zone (Morrice et al., 1997). In contrast, the relative transient storage zone size (AS/A = 0.24, cross-sectional area of the main channel A 3.0 m2, cross-sectional area of storage zone AS 0.7 m2) determined as median value of five sampling campaigns in March, May, July, September and December 2002 with OTIS-P was relatively large (Gücker and Pusch, 2006). Therefore, from a hydrological point of view the hyporheic zone has the potential to substantially contribute to micropollutant attenuation. A central prerequisite for the analytical solution of the transport equation used to calculate seepage flows is that the major flow direction in the streambed is vertical. This is a justified simplification since the hydraulic conductivity of the fine muddy streambed sediment is lower than the hydraulic conductivity of the underlying aquifer material. Thus, the streambed sediment is the strongest hydraulic resistance in surface water–groundwater exchange. Water crossing the hydraulic resistance prefers the easiest flow path. In case of a homogenous, isotropic streambed the shortest flow path is the easiest path, i.e. vertical flow is a good approximation for the flow direction. The two sites of the present study differ substantially between each other in the exchange between surface water and sediment. Based on the temperature profiles, we determined a substantial infiltration of surface water into the hyporheic zone at site 1, while at site 2 groundwater exfiltrated into the river at a very low rate. Irregularities in the concentration profiles (Figs. 3 and 4) might be attributed to the fact that there are also horizontal flow components although this is not indicated by the temperature profiles (Fig. 2). Interpretation of chemical data is complicated due to temporal changes of the surface water composition. WWTP effluents show daily and weekly fluctuations (both of discharge and loads), which are superimposed by precipitation events, and thus the upper (concentration) boundary condition is variable. Moreover, the pore water, sampled at the different ports of a nested piezometer, has travelled different distances in the streambed and infiltrated at different times. Thus, we cannot directly compare the concentrations measured in the different depths since the initial concentrations of the water that infiltrated into the streambed might have been different. Thus— assuming that borate is conservative in the sediment (Vengosh et al., 1999)—the increasing borate concentration with depth (site 1; Fig. 3) might reflect such different concentrations in the surface water. Based

on the temperature profiles and a porosity of 0.2 we estimated a downward directed flow velocity of approximately 20 cm d−1, i. e. the water in 5 cm depth (first sampling port) infiltrated approximately 6 h before sampling the surface water at the site, the water in 15 cm (second sampling port) infiltrated 18 h earlier and water in 95 cm depth (lowest sampling port) infiltrated about 5 days earlier. Thus, the surface water composition during the 5 days prior to pore water sampling has impacts on the subsurface micropollutant concentrations. In addition, we can conclude from travel times and low flow velocities that there is much time for potential micropollutant transformation in the hyporheic zone. Generally, the observation that surface water infiltrates at site 1, derived from the temperature profiles, is confirmed by borate since its concentrations are substantially above the background level. For site 2 the temperature profiles indicate a minor exfiltration, which is also supported by the borate profiles. There, concentrations of 0.1 mg L−1 throughout the profile are interpreted as background concentration in the anthropogenicallyimpacted aquifer. The concentration of non-geogenic Gd in the streambed is also in agreement with this interpretation. However, the profiles of borate and non-geogenic Gd profiles are quite different at site 1 (Fig. 3); as mentioned above these differences might result from varying surface water composition and different times of infiltration of surface water into the hyporheic zone. 4.2. Biogeochemistry Due to the high content of organic matter in the upper soil layers of floodplains combined with water levels close to the surface, reducing conditions in floodplain aquifers occur (Lewandowski and Nützmann, 2010). At site 1 (A, B, C) shortly behind the WWTP outlet fine particulate material (including iron(oxy)hydroxides with sorbed phosphorus) and easily degradable organic matter settle onto the streambed. Electron acceptors like oxygen and nitrate are consumed within the first millimetres of the streambed sediment resulting in nitrate concentrations below the detection limit throughout the streambed profile (Fig. 4). Easily degradable organic matter is continuously delivered from the overlying water. Therefore, the most intense mineralisation occurs close to the streambed surface. The content of organic matter (loss on ignition, carbon content) is higher at site 1 (A, B, C) than at site 2 (D, F) due to its vicinity to the discharge of the treated wastewater. The concentration of ammonium—originating from the mineralization of organic matter—drastically increases in the uppermost streambed layer compared to the overlying water while only little further increase occurs with depth. The concentrations of dissolved iron and dissolved manganese also increase in the uppermost streambed sediment layers due to reductive dissolution of oxidized iron and manganese compounds. Sulphate concentrations in the infiltrating surface water decrease drastically within the uppermost streambed layers after consumption of all other, more favourable electron acceptors, i.e. strictly anoxic conditions prevail in the streambed sediments (A, B). The sulphate profile at site C—close to the left bank—decreases much less with depth. The most probable explanation is that heterogeneity in the streambed sediment might have caused a preferential flow path bypassing the most reductive sediment layers or at least crossing it faster than the sulphate reduction requires for complete sulphate consumption. This would also explain why the non-geogenic Gd profile C differed from profiles A and B. In deeper streambed layers decreasing concentrations of dissolved iron and dissolved manganese are caused by less reductive conditions and an import of small amounts of oxygen from macrophyte roots (Hupfer and Dollan, 2003). Dissolved total phosphorus concentrations in the stream Erpe are high and increase drastically in the streambed sediments due to mineralisation of organic matter containing phosphorus (Hendricks and White, 2000) and due to reductive dissolution of iron- and manganese compounds releasing sorbed phosphorus (Vanek, 1991; Surridge et al., 2005).

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At site 2 we measured exfiltration of groundwater at a very low rate. From a nearby site at another river we know that nitrate is efficiently denitrified shortly after groundwater enters the floodplain (Lewandowski and Nützmann, 2010). Analogously, the absence of nitrate is assumed for the floodplain of the stream Erpe. Nitrate concentrations in the hyporheic zone (sampling ports from 5 to 95 cm in the streambed sediment) are below the detection limit. Ammonium concentrations in the exfiltrating groundwater are high as a consequence of the reducing redox milieu and mineralisation processes occurring in the aquifer of the floodplain. Sulphate concentrations in the aquifer of the floodplain are low (about 20 mg L−1). Close to the sediment surface the redox potential is even lower, due to the settling of easily degradable organic matter onto the streambed surface. Therefore, sulfate concentrations decrease even further close to the sediment surface. At site 2 iron dissolved in the groundwater with its low redox potential it is transported upwards. In the immediate vicinity of the sediment– water interface the oxidation of dissolved iron(II) results in precipitation of iron(III)-compounds. Different processes (gas ebullition, bioturbation, current-induced transport of bed sediment) result in a mixing of the uppermost sediment layers (Meysman et al., 2006; Sondergaard et al., 1993; Martens and Val Klump, 1980). Iron(III)compounds mixed into deeper sediment layers are gradually reduced due to the low redox potential occurring there (Ferro et al., 2003; Vanek, 1991). Consequently, iron(II) concentrations increase close to the surface. The intensity of the processes involved in iron turnover differs between D and F. This is caused by heterogeneities of bed sediments or different compositions of the groundwater exfiltrating at both river banks (further parameters not shown here like calcium, magnesium, sodium, potassium also indicate different water compositions). In principle, the same interpretation applies also for manganese. Dissolved total phosphorus in the exfiltrating groundwater is relatively low for groundwater underlying a floodplain. Due to mineralization of organic matter and the reductive dissolution of iron compounds, some increase of phosphorus concentrations occurs close to the sediment–water interface. 4.3. Micropollutants Borate and non-geogenic Gd indicated a minor dilution of the stream water from site 1 to site 2 and thus we can directly compare micropollutant concentrations of both sites. A small loss of bezafibrate and indomethacin and a clear loss of ketoprofen (concentration at site 2 approx. 78% lower) were detected within the distance of 2 km between both sites (Fig. 5). This reduction might indicate attenuation of ketoprofen in the surface water (aerobic degradation or photolysis) or during the temporary storage of surface water in the hyporheic zone (sorption/biodegradation). However, since the magnitude of the ketoprofen reduction was much higher than we would attribute to such processes based on literature data (Daneshvar et al., 2010; Radke et al., 2010), with our current knowledge we assume that varying concentrations in the effluent of the WWTP caused the observed differences (compare 4.1). Regarding the attenuation of the micropollutants, the hyporheic zone at site 2 (D, F) is of minor importance. As described above, we inferred from temperature, borate and non-geogenic Gd depth profiles a slight exfiltration of groundwater into the stream Erpe. Thus, no infiltration of surface water with its high micropollutant load into the hyporheic zone is assumed except for the upper 20 cm of piezometer F (compare borate and non-geogenic Gd profiles). The profiles of micropollutant concentrations measured at site 2 (D, F) are in general agreement with the observation that exfiltration and no infiltration occurs at site 2 (Fig. 5). To evaluate the attenuation of micropollutants in the hyporheic zone, the depth profiles of the different micropollutants at site 1 (A, B, C) are more insightful than those at site 2. Compared to the surface

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water, the concentrations of all micropollutants except clofibric acid were lower in the hyporheic zone. However, the concentrations of borate and non-geogenic Gd, which are regarded as conservative indicators for the presence of surface water, also decreased with depth. The lower concentrations result at least partly from a higher dilution of the compounds originating from treated wastewater, i.e. concentrations of all compounds in the surface water were lower at the time the hyporheic water entered the hyporheic zone. To evaluate micropollutant attenuation in the hyporheic zone, it is, therefore, necessary to normalize the micropollutant concentrations to borate or non-geogenic Gd (Figs. 3, 6). Unfortunately, the profiles of borate and non-geogenic Gd (Fig. 3) are quite different. If both were perfect sewage indicators their profiles should be similar. Fig. 6 shows four alternative normalization strategies: micropollutant concentrations are normalized to surface water concentrations or to pore water concentration in the uppermost streambed layer (0– 10 cm). Then they are related to borate profiles or non-geogenic Gd profiles to account for different percentages of surface water in the different samples. All normalized profiles are based on the assumption of constant proportions of the micropollutants and the sewage indicators borate and non-geogenic Gd in the effluent of the WWTP. Normalized ratios above 1 in Fig. 6 indicate that this assumption is neither valid for profiles related to non-geogenic Gd nor for profiles related to borate. We conclude that there must be significant temporal changes in the composition of the effluent of the WWTP (compare 4.1) such as those reported for other catchments (Pailler et al., 2009). Keeping the limitations of the normalization to the conservative wastewater indicators in mind, the most logical results were obtained by normalising the micropollutant concentrations to the uppermost sediment layer and in relation to borate concentrations (Fig. 6D). Naproxen and indomethacin are not discussed since their subsurface concentrations are quite low resulting in a large error of normalized values. For all other compounds except bezafibrate, the normalized concentration ratios reflect a substantial decrease with depth. From these ratios, we can deduce that bezafibrate reflects a compound which is not attenuated in the first 50 cm of the profile, while we observed an increasing attenuation of diclofenac, ibuprofen, ketoprofen, and clofibric acid with increasing depth and thus increasing retention time in the hyporheic zone. Since dilution is eliminated by the normalization to borate, this attenuation can be attributed to sorption and biotransformation processes. Based on the comparison of the five compounds, the attenuation rate of the micropollutants in the hyporheic zone is in the order bezafibrate b diclofenac b ketoprofen b ibuprofen ≈ clofibric acid. The same attenuation rate of clofibric acid and ibuprofen is surprising. In previous studies of WWTPs and the behaviour of the compounds in surface waters and sediments, ibuprofen has been reported to be easily biodegradable under aerobic conditions while clofibric acid was usually persistent under these conditions; the sorption properties of both compounds are similar. However, little is known about the fate of the investigated micropollutants in extremely anaerobic systems such as the hyporheic zone at site 1. If biodegradation was the dominant attenuation mechanism in the studied system, we would expect clofibric acid to be the compound with the lowest and ibuprofen the one with the highest attenuation rate. Since this is not the case, we hypothesize that the observed attenuation of the micropollutants in the studied hyporheic zone is more likely caused by sorption rather than by transformation processes or by a combination of both. However, previous studies have shown that these pharmaceuticals are only weakly adsorbing to soils (Xu et al., 2009), aquifer material (Scheytt et al., 2004), and river sediments (Kunkel and Radke, 2008) with solid-water partitioning coefficients (Kd) lower than 20 L kg−1 (Xu et al., 2009; Scheytt et al., 2004). This can be attributed to the acidic nature of the studied pharmaceuticals. As the dissociation constant (pKa) of all

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compounds is below 4.5 (see Table 1), their anionic species is dominating at the prevailing pH in most aquatic systems. As sorption of anions to environmental surfaces is generally little compared to neutral or cationic species, we assume that the high sedimentation in the stream Erpe due to the WWTP effluent continuously generates new sorption sites for micropollutants at the immediate sediment– water interface. However, given the uncertainties of this preliminary study (mainly varying composition of WWTP effluent, substantial differences between the depth profiles of the two wastewater indicators) this interpretation is still somewhat speculative. 5. Conclusions and outlook The results of our preliminary study show that the hydrology of a site has a major impact on redox milieu and micropollutant concentrations in the hyporheic zone. At site 2 high concentrations of organic micropollutants occur in the investigated streambed at least down to the maximal studied depth (1 m). Several micropollutants showed decreasing concentrations with depth, but since the concentrations of sewage indicators (non-geogenic Gd, borate) also decreased with depth we cannot conclude from decreasing micropollutant concentrations that attenuation of micropollutants occurs during passage of the hyporheic zone. The major drawback of the present study is the varying composition of the WWTP effluent, which inhibits a consistent interpretation of the micropollutant concentrations. In a future study surface water should be sampled continuously as well as at a sufficient temporal resolution to determine the temporal course of the concentrations of sewage indicators (borate and non-geogenic Gd) and micropollutants. This would allow estimating minimal, maximal and average concentrations. Furthermore, the measurement of boron isotopes instead of borate might be helpful to distinguish between different sources of borate contamination and distinguish more clearly between background borate concentrations and effluent of the WWTP Münchehofe. In future studies the temporal course of micropollutant concentrations in the streambed (nested piezometers) in combination with a frequent sampling of surface water concentrations might be useful to learn more about the fate of micropollutants in the hyporheic zone. Additional sites with different hydrologic regime should also be equipped with nested piezometers. For a first overview about the hydrologic regime along the stream distributed temperature sensing might be useful. Furthermore, the strictly vertical flow direction assumed in the present sampling campaign should be verified by direct measurements, e.g. with a heat pulse sensor (Lewandowski et al. submitted). Knowledge of flow paths in the hyporheic zone is essential for a sound and comprehensive data interpretation since the optimum sampling strategy would be to sample along individual flow paths. However, at present we are lacking appropriate techniques and methods to directly measure such flow paths in the hyporheic zone. The differences of the micropollutant concentrations in the surface water between sites 1 and 2 integrate all processes occurring within the 2 km of the stream Erpe. The only small changes of micropollutant concentrations along this stretch could be due to two reasons: (1) There are only minor attenuation processes in the surface water and in the hyporheic zone or (2) there is—on a quantitative basis—little exchange between the surface water and the hyporheic zone. A more detailed sampling strategy (see above) and an estimation of exchange rates of surface and subsurface water will be required to elucidate the correct explanation. Unfortunately, based on the current data we cannot test the hypothesis that the hyporheic zone is a compartment of utmost importance for the transformation of micropollutants in lowland rivers. The present study showed that the approach of nested piezometers is useful for studying micropollutant distribution throughout the hyporheic zone, but many more extensive studies

are required to adequately test the aforementioned hypothesis. Furthermore, the present study could not clarify whether nongeogenic Gd or borate or a combination of both parameters is the best sewage indicator. This study clearly showed the close coupling of hydrology, biogeochemistry and micropollutant concentrations in the hyporheic zone. The investigation of the relevant processes and their interactions requires an interdisciplinary approach and specific investigations that should be addressed in detail in future studies. Acknowledgements This study was conducted as a pre-investigation for the research proposal Interhyp. We thank all persons involved in this proposal for inspiring the present research. Furthermore, Christine Sturm (IGB), Hans-Jürgen Exner (IGB), Elke Zwirnmann (IGB), Hella Schmeisser (TU Berlin), Jutta Jakobs (TU Berlin) and Katrin Noak (TU Berlin) are acknowledged for their help during field work and analytics. We thank Björn Gücker (Federal University of São João del-Rei, Minas Gerais, Brazil) for providing us with data previously collected at the study site. Stefanie Burkert (IGB) is acknowledged for editing the manuscript. Two anonymous reviewers and the editor Ian Snape are acknowledged for checking the manuscript. References Al-Ahmad A, Daschner FD, Kümmerer K. Biodegradability of cefotiam, ciprofloxacin, meropenem, penicillin g, and sulfamethoxazole and inhibition of waste water bacteria. Arch Environ Contam Toxicol 1999;2:158–63. Bau M, Dulski P. Anthropogenic origin of positive gadolinium anomalies in river waters. Earth Planet Sci Lett 1996;143:245–55. Birgand F, Skaggs RW, Chescheir GM, Gilliam JW. Nitrogen removal in streams of agricultural catchments—a literature review. Crit Rev Environ Sci Tech 2007;37:381–487. Bredehoeft JD, Papadopolus IS. Rates of vertical groundwater movement estimated from earth's thermal profile. Water Resour Res 1965;2:325–8. Cardenas MB. Dynamics of fluids, heat and solutes along sediment–water interfaces: a multiphysics modeling study. Socorro: New Mexico Institute of Mining and Technology; 2006. Daneshvar A, Svanfelt J, Kronberg L, Weyhenmeyer G. Winter accumulation of acidic pharmaceuticals in a Swedish river. Environ Sci Pollut Res 2010;17:908–16. Daughton CG. Cradle-to-cradle stewardship of drugs for minimizing their environmental disposition while promoting human health. I. Rationale for and avenues toward a green pharmacy. Environ Health Perspect 2003;111:757–74. Elbaz-Poulichet F, Seidel J, Othoniel C. Occurrence of an anthropogenic gadolinium anomaly in river and coastal waters of southern France. Water Res 2002;36:1102–5. Ferro PI, Van Nugteren J, Middelburg J, Herman PMJ, Heip CHR. Effect of macrofauna, oxygen exchange and particle reworking on iron and manganese sediment biogeochemistry: a laboratory experiment. Vie Mileu Environ 2003;53:211–20. Gruenheid S, Huebner U, Jekel M. Impact of temperature on biodegradation of bulk and trace organics during soil passage in an indirect reuse system. Water Sci Technol 2008;57:987–94. Gücker B, Pusch MT. Regulation of nutrient uptake in eutrophic lowland streams. Limnol Oceanogr 2006;51:1443–53. Gücker B, Brauns M, Pusch MT. Effects of wastewater treatment plant discharge on ecosystem structure and function of lowland streams. J N Am Benthol Soc 2006;25:313–29. Haggard BE, Storm DE, Stanley EH. Effect of a point source input on stream nutrient retention. J Am Water Resour Assoc 2001;37:1291–9. Harding AP, Wedge DC, Popelier PLA. pKa prediction from “Quantum chemical topology” Descriptors. J Chem Inf Model 2009;49:1914–24. Hendricks SP, White DS. Stream and groundwater influences on phosphorus biogeochemistry. In: Jones JB, Mulholland PJ, editors. Streams and ground waters, aquatic ecology series. San Diego: Academic; 2000. p. 221–35. Heppell CM, Wharton G, Cotton JAC, Bass JAB, Roberts SE. Sediment storage in the shallow hyporheic of lowland vegetated river reaches. Hydrol Process 2009;23: 2239–51. Huber MM, Canonica S, Park GY, von Gunten U. Oxidation of pharmaceuticals during ozonation and advanced oxidation processes. Environ Sci Technol 2003;37:1016–24. Hupfer M, Dollan A. Immobilisation of phosphorus by iron-coated roots of submerged macrophytes. Hydrobiologia 2003;506(1–3):635–40. House WA. Geochemical cycling of phosphorous in rivers. Appl Geochem 2003;18: 739–48. Hynes HBN. The biology of polluted waters. Toronto, Ontario: University of Toronto Press; 1974. Kalsch W. Biodegradation of the iodinated X-ray contrast media diatrizoate and iopromide. Sci Total Environ 1999;225:143–53. Köhler J, Gelbrecht J, Pusch M. Die Spree–Zustand, Probleme, Entwicklungsmöglichkeiten. Stuttgart: Schweizbart; 2002. Krause S, Hannah DM, Fleckenstein JH. Hyporheic hydrology: interactions at the groundwater–surface water interface Preface. Hydrol Process 2009;23:2103–7.

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