Effects of municipal wastewater on aquatic ecosystem structure and function in the receiving stream

Science of the Total Environment 454–455 (2013) 401–410

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Science of the Total Environment journal homepage: www.elsevier.com/locate/scitotenv

Effects of municipal wastewater on aquatic ecosystem structure and function in the receiving stream Dominic Englert a,⁎, Jochen P. Zubrod a, Ralf Schulz a, Mirco Bundschuh a, b a b

Institute for Environmental Sciences, University of Koblenz-Landau, Fortstrasse 7, D-76829 Landau, Germany Department of Aquatic Sciences and Assessment, Swedish University of Agricultural Sciences, Uppsala, Sweden

H I G H L I G H T S • Wastewater affects structural and functional endpoints in ecosystem. • Adverse effects occur some hundred meters below the effluent if dilution potential is low. • Powdered activated carbon effectively reduces ecotoxicity of wastewater.

a r t i c l e

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Article history: Received 13 January 2013 Received in revised form 4 March 2013 Accepted 6 March 2013 Available online 9 April 2013 Keywords: Ecosystem functions Leaf decomposition In situ Micropollutants Gammarus Powdered activated carbon

a b s t r a c t During recent years, increasing incidences of summer droughts – likely driven by climate change – reduced the dilution potential of low-order streams for secondary treated wastewater also in temperate Europe. Despite the potential risks to ecosystem integrity, there is a paucity of knowledge regarding the effects of different wastewater dilution potentials on ecosystem functions. The present study investigated the implications of secondary treated wastewater released into a third-order stream (Queich, southwest Germany) during a season with low dilution potential (summer; ~ 90% wastewater) as compared to a season with high dilution potential (winter; ~ 35% wastewater) in terms of leaf litter decomposition and macroinvertebrate communities. Adverse effects in macroinvertebrate mediated leaf mass loss (~65%), gammarids' feeding rate (~ 80%), leaf associated fungal biomass (>40%) and shifts in macroinvertebrate community structure were apparent up to 100 and 300 m (partially 500 m) downstream of the wastewater treatment plant effluent during winter and summer, respectively. In addition, a Gammarus fossarum laboratory feeding trial demonstrated the potential of powdered activated carbon to reduce the ecotoxicity of released wastewater. These results urge the development and evaluation of adequate management strategies, e.g. the application of advanced wastewater treatment technologies, to protect the integrity of freshwater ecosystems, which is required by the European Water Framework Directive — also considering decreasing dilution potential of streams as projected by climate change scenarios. © 2013 Elsevier B.V. All rights reserved.

1. Introduction Many chemicals of anthropogenic origin are only partly retained or degraded in municipal wastewater treatment plants (WWTPs) by conventional secondary treatment (i.e. mechanical and biological). Therefore, WWTP effluents are considered as one of the major sources of nutrients (e.g. Martí et al., 2010) as well as inorganic (Dyer and Wang, 2002) and organic (micro)pollutants (Bueno et al., 2012; Daughton Abbreviations: CI, confidence interval; NMDS, non-metric multidimensional scaling; PAC, powdered activated carbon; PERMANOVA, permutational multivariate analysis of variance; SIMPER, similarity percentage analysis; WWTP, wastewater treatment plant. ⁎ Corresponding author at: Institute for Environmental Sciences, University of Koblenz-Landau, Landau Campus, Fortstrasse 7, 76829 Landau, Germany. Tel.: +49 6341 280 31 328. E-mail address: [email protected] (D. Englert). 0048-9697/$ – see front matter © 2013 Elsevier B.V. All rights reserved. http://dx.doi.org/10.1016/j.scitotenv.2013.03.025

and Ternes, 1999) for aquatic ecosystems (Schwarzenbach et al., 2006). The structure of aquatic communities inhabiting these ecosystems may suffer from this continuous contamination directly (Bundschuh et al., 2011a) or indirectly (Bundschuh et al., 2011b; Duddridge and Wainwright, 1980). In addition, ecosystem functions including the decomposition of leaf litter, which provides energy in the form of fine particulate and dissolved organic matter for local as well as downstream communities (Cummins and Klug, 1979), may be affected by conventionally treated wastewater (Bundschuh et al., 2011a). Implications of wastewater on both ecosystem structure and function were, however, until now, mainly assessed in effluent dominated streams of semi-arid regions (e.g. Canobbio et al., 2009; Ortiz et al., 2005; Rueda et al., 2002; Ruggiero et al., 2006). Although previous studies indicated either no or only negligible impairments of secondary treated wastewater in the ecosystem function of leaf litter decomposition in temperate

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regions (Gücker et al., 2006; Ladewig et al., 2006; Spänhoff et al., 2007; Suberkropp et al., 1988), the increasing frequency and intensity of summer droughts, which are potentially due to climate change, have resulted in decreased water levels, particularly in low-order streams, thereby additionally reducing their dilution potential (Prellberg, 2010). Therefore, it may be hypothesized that Central European regions with low precipitation (i.e. Palatinate, Germany; Ramm, 2005), already suffer from a low dilution rate of secondary treated wastewater released into streams, especially during the dry summer months. However, there is a scarcity of knowledge regarding the implications on ecosystem functions, such as the leaf litter decomposition, during seasons with low (summer) compared to high (winter) dilution potential within the receiving stream (cp. Bundschuh et al., 2011c). Knowledge in this field may help to predict future implications of water scarcity and to develop appropriate management strategies. To address this gap in scientific knowledge, the present study assessed the potential adverse effects of secondary treated wastewater released from the WWTP Landau-Mörlheim, Germany, on leaf litter decomposition. The study was conducted within a third-order stream, the Queich, during both winter and summer seasons assuming that effects are more pronounced during low flow conditions (= summer). Alterations in the macroinvertebrate community were assessed by surber sampling as well as by collecting organisms associated with coarse-mesh leaf litter bags. Moreover, bacterial cell numbers and fungal biomass were assessed to enable the determination of changes in the activity of leaf-associated microbial decomposers, bacteria and fungi, which may trigger the changes seen in leaf litter decomposition process, such as leaf mass loss. In addition, in situ bioassays investigating the feeding rate of the leaf-shredding amphipod Gammarus fossarum were performed. Such in situ bioassays are supposed to display exclusively direct – not food quality related – implications as discussed by Bundschuh and Schulz (2011) for a comparable laboratory-based experimental set-up. In order to achieve regulatory requirements such as the European Water Framework Directive (European Commission, 2000) also during seasons with low dilution potential, advanced treatment technologies may be useful in the medium term (Stalter et al., 2010), which is, however, insufficiently underpinned from an ecotoxicological viewpoint (Bundschuh et al., 2011d). Powdered activated carbon (PAC) is one technology with the potential to reduce the load of (in)organic micropollutants in wastewater. Therefore, the present study bolstered the (semi)field experiments by a laboratory feeding trial – also using G. fossarum – investigating the potential of PAC for the reduction of wastewaters' ecotoxicity.

2.2. Experimental design For the present study, the control site (LD1w/s) was located in the Queich's main stream 250 m upstream of the WWTP effluent (Fig. 1). This site is not affected by WWTPs as no effluent was located up to 15 km further upstream. To assess the ecotoxicological implications of secondary treated wastewater released into the Queich, macroinvertebrate- and microorganism-mediated leaf mass loss, leaf associated microbial endpoints (bacterial cell number and fungal biomass) as well as shifts in the macroinvertebrate community structure were assessed 100 m (LD2w/s) downstream during winter (w; 2010/11 for six weeks after 14, 28 and 42 d of exposure) by using the methods described below. As substantial effects in all endpoints were observed during winter (see results section), additional sampling sites (up to 1000 m further downstream) were assessed during the experiments conducted in the following summer season to determine the area downstream that was affected by the wastewater. Hence, during the summer (s; 2011 only three weeks due to a faster decomposition as detailed below) three additional sampling sites located 300 (LD3s), 500 (LD4s) and 1000 m (LD5s) downstream of the WWTP effluent were investigated (Fig. 1). 2.3. Wastewater dilution and water parameters Water quality parameters were measured in situ weekly at all sites during winter and summer. Oxygen saturation, temperature, pH and conductivity were measured with a WTW Multi 340i/SET (Wissenschaftlich Technische Werkstätten GmbH, Weilheim, Germany). Current velocity was measured with a Höntzsch instrumentals flow meter (type μO-TAD; Waiblingen, Germany) respectively. Additional parameters including ammonium, nitrate, nitrite, phosphate, chloride, sulfate concentrations and hardness were quantified using MachereyNagel visocolor® kits (Macherey-Nagel, Düren, Germany) in the laboratory. Moreover, the seasonal contribution of wastewater to the branch's discharge was calculated during both seasons. For this, the volumetric flow rate (w: 467.0 L/s; s: 28.8 L/s) of the Queich's branch was calculated by multiplying the stream cross-sectional area (upstream of the WWTP effluent; 2.44 and 0.24 m2) with the mean flow velocity (of the crosssection; 0.20 and 0.12 m/s) both measured during winter and summer, respectively. The average contribution of the wastewater to the discharge of the Queich's branch was calculated by considering 250 L/s as the average WWTP discharge (EWL, 2012) for both seasons. 2.4. Macroinvertebrate sampling

2. Materials & methods 2.1. Study site The present study was performed in a third-order stream (the Queich) within the region of Palatinate, Germany near the city of Landau. The municipal WWTP Landau-Mörlheim (49°12′N; 8°10′E) releases wastewater of a population equivalent of 80,000, 30,000 of which comes from industry (Raisin, EWL, personal communication), into the Queich. The secondary treated wastewater (200–300 L/s; EWL, 2012) is initially discharged through a sewage channel without aquatic macrophytes, which enters after approximately 2 km into a branch of the Queich. Since no dilution takes place within the sewer, this inflow point is referred to as “WWTP effluent” in the remainder of the manuscript. Approximately 350 m downstream of the WWTP effluent, the branch enters the Queichs' main stream, further diluting the introduced wastewater. The sandy streambed of the investigated Queich section had a width of 3–4 m and a water depth between 0.5 to 0.7 m (up to 1.5 m in pools that were scattered along the area), which varied among sites and seasons. Trees of the genera Alnus, Acer, Prunus and Sambucus dominated the riparian vegetation along the investigated stream section.

A Surber Sampler (500-μm mesh-size; surface: 1/8 m2; Surber, 1970) was used to obtain a representative assessment of the macroinvertebrate community composition (Storey et al., 1991). Three samples were taken at the beginning as well as at the termination of the experiments conducted in the winter season. Since, during the winter, no significant differences were observed regarding the community composition at both sampling sites between sampling dates (see Supplemental data Fig. A1), macroinvertebrates were sampled only once in summer; this occurred after ten days of the beginning of the study (i.e. in the middle of the experiment). Organisms were fixed in 70% ethanol, counted and identified to the lowest possible taxonomic level (genus for the majority of taxa; Table 1). 2.5. Leaf litter bags Leaves of Alnus glutinosa (L.) Gaertn. (black alder), a common species in riparian zones of temperate Europe (Hewitt, 1999) and present at the investigated stream section, were utilized to assess leaf litter decomposition in terms of leaf mass loss at each sampling site. Leaves were collected shortly before defoliation in October 2008 from trees near Landau, Germany (49°11´N; 8°05´E), and were stored

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Fig. 1. Location of WWTP Landau-Mörlheim as well as the sampling sites within the Queich investigated during winter 2010/11 (w) and summer 2011 (s); m-values indicate distance downstream of the WWTP effluent.

frozen (− 20 °C) until further use. For the leaf litter decomposition studies conducted in the winter, 2 (± 0.1) g of leaves dried at 60 °C, pre-weighed and re-soaked, were placed in coarse-mesh leaf litter bags (mesh size: 10.0 mm; five bags per site and retrieval date), allowing the stream macroinvertebrate community to feed on the material. An equal amount of black alder leaves was introduced into finemesh leaf litter bags (type I; mesh size: 0.5 mm; five bags per site and retrieval date) accounting for leaf mass loss caused by microbial degradation and leaching. Because faster leaf litter decomposition was hypothesized to occur during the summer (e.g. Seastedt et al., 1983), it was necessary to introduce 3 (±0.1) g of leaf litter into coarse-mesh leaf litter bags to insure the provision of sufficient amounts of leaf material for the leaf shredding macroinvertebrates during the investigation period. To avoid a bias in the quantification of leaf mass loss caused by microbial activity and leaching, eight black alder leaves were used for microbial analysis (i.e. fungal biomass and bacterial abundance). These leaves were treated in the same manner as those used for the leaf litter decomposition experiments and were allocated to each sampling site in winter as well as summer using five additional fine-mesh leaf litter bags (type II) per site and retrieval date. Five bags from each type (i.e. coarse, fine type I and II) were retrieved after 14, 28 and 42 d of exposure during winter. Due to the increased number of sampling sites investigated during summer, compared to the winter season, as well as the expected faster leaf decomposition (detailed above), leaf bags were retrieved after 21 d of exposure. The remaining leaf litter in each bag was gently rinsed with tap water and macroinvertebrates associated with coarse-mesh leaf litter bags, referred to as bycatch in the remainder of the manuscript, were washed through a 500-μm sieve. Leaf litter from coarseand fine-mesh leaf litter bags (type I) were dried (60 °C, 48 h) and weighed to the nearest mg. The bycatch were identified, sorted and counted at the order level (i.e., Amphipoda, Ephemeroptera, Isopoda, Plecoptera and Trichoptera) to obtain further information about the leaf associated macroinvertebrates at each study site. Finally, the biomass of macroinvertebrate orders was determined as dry-weight (60 °C, 48 h) to the nearest 0.01 mg (Supplemental data Table A1). 2.6. Leaf-associated microbial parameters Three out of five fine-mesh leaf litter bags (type II) per site were randomly chosen to analyze bacterial cell numbers as well as fungal biomass at the termination of the experiment during winter and summer, respectively (differences in the colonization dynamics among study sites were beyond the scope of the present study). Ergosterol, a component of eumycotic cell walls (Gessner and Schmitt, 1996), was used as a proxy for fungal biomass. Ergosterol was extracted in alkaline

methanol from freeze-dried leaf material. Following purification by solid-phase extraction (Sep-Pak® Vac RC tC18 500 mg sorbent; Waters, Milford, USA), ergosterol was quantified by high-performance liquid chromatography (HPLC; 1200 Series, Agilent Technologies, Santa Clara, USA) measuring absorbance at 282 nm. Ergosterol concentrations were finally converted to fungal biomass assuming an average mycelia concentration of 5.5 mg ergosterol/g fungal dry weight (Gessner and Chauvet, 1993). Bacterial cell numbers were determined by epifluorescence microscopy (Buesing, 2005). In brief, formaline-preserved cells were detached from the leaf discs via ultrasonication. A 5-μL aliquot of each sample was filtered using an aluminum oxide filter (pore size 0.2 μm, Anodisc, Whatman, Beckenham, UK) and the bacterial cells were dyed using SYBRGreen II (Molecular Probes®, Eugene, USA). The cells were counted on digital photographs of at least ten microscopic fields per replicate using an image analysis system (Axio-Vision Rel. 4.8, Carl Zeiss MicroImaging, Jena, Germany) and were normalized to the dry weight of the leaf discs. 2.7. In situ bioassays and laboratory feeding trial In situ bioassays were conducted to assess direct implications of wastewater on the dominant shredder genus (i.e. Gammarus) within the stream by measuring its feeding rate on standardized leaf material (i.e. leaf discs). Leaf discs were prepared according to Bundschuh et al. (2011c). In brief, discs of 2.0 cm diameter were cut from frozen (−20 °C) black alder leaves (collected shortly before defoliation in October 2008; see 2.5). Leaf discs were subsequently conditioned in a nutrient medium (Dang et al., 2005) for 10 d in combination with black alder leaves previously exposed in the near natural stream Rodenbach, Germany (49°33′N, 8°02′E) in order to establish a natural microbial community consisting of fungi and bacteria. These microorganisms alter the leaves (physically and chemically) and, thereby, modify their palatability and nutritional value for leaf shredding organisms (Bärlocher, 1985). After conditioning, the leaf discs were dried at 60 °C and weighed to the nearest 0.01 mg to ensure an accurate measurement of the amphipods' feeding rate (Maltby et al., 2002). Finally, leaf discs were re-soaked in tap water and test medium for 48 h prior to the start of each in situ bioassay and the laboratory feeding trial respectively. G. fossarum was chosen as the test species for the in situ bioassays as well as the laboratory feeding trial, since it naturally occurs with other gammarids at the study site (Table 1). The test organisms, however, were collected from the Hainbach, a near natural stream, close to Landau, Germany (49°14´N; 8°03´E), 24 h prior to the start of each in situ bioassay but one week before the initiation of the laboratory feeding trial to ensure acclimation to laboratory conditions.

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Table 1 Mean (±SE) abundance of macroinvertebrates obtained by surber sampling (surface: 1/8 m2) during winter (n = 6) and summer (n = 3). Winter

Summer

LD1w Aphelocheirus aestivalis Asellus aquaticus Baetis sp. Oligoplectrum maculatum Calopteryx sp. Chironomidae Corbicula sp. Elmis sp. Ephemerella sp. Gammarus fossarum Gammarus pulex Gammarus roeseli Gomphus vulgatissimus Ecdyonurus sp. Electrogena affinis Hirudinea Hydraenidae Hydropsyche sp. Lepidostoma hirtum Lestes barbarus Limnephilidae Lymnaeidae Odontoceridae Physella acuta Rhyacophilidae Sphaerium sp. Tabanidae Dicranota sp. Tubifex sp. Simulidae Mean abundance

1 1 2 – – 4 – – – 23 30 3 – – 1 1 1 1 1 – 1 – – – – 1 – 1 2 – 70

LD2w (±1) (±0) (±1)

(±3)

(±11) (±16) (±2)

(±0) (±0) (±0) (±0) (±0) (±1)

(±0) (±0) (±1) (±15)

– 2 – – – 4 1 1 – 4 1 – – – – – – – – – 1 – – – 1 1 1 1 – 1 14

LD1s (±1)

(±3) (±0) (±0) (±3) (±1)

(±1)

(±0) (±0) (±0) (±0) (±0) (±7)

2 – – – – 40 – – – 20 33 – – 1 – – – 1 – – 3 1 2 – – – – 1 57 – 159

LD2s (±1)

(±36)

(±15) (±20)

(±0)

(±0)

(±3) (±1) (±2)

(±0) (±33) (±42)

During this time, animals for the in situ bioassays and the laboratory feeding trial were fed with conditioned black alder leaves ad libitum and kept at 20 ± 1 °C in aerated stream water and test medium (SAM-S5; Borgmann, 1996), respectively. Specimens were visually checked for acanthocephalan parasites prior to use, since parasites may affect amongst others the feeding behavior of gammarids (Pascoe et al., 1995) and, hence, the endpoint (i.e. feeding rate) investigated during the present experiments. Only un-parasitized adult animals of approximately 4.5 to 6.0 mm in length (Pöckl, 1992) were used. In situ bioassays (each 7 days long) were conducted during six and three consecutive weeks – parallel to the leaf litter bag studies – during the winter and summer season, respectively, as described in Bundschuh et al. (2011a). In brief, one individual of G. fossarum was placed with two pre-weighed conditioned leaf discs in a cage covered by a 1.0 mm mesh screen. Twenty of those cages were exposed at each site. Five additional cages without G. fossarum were set up to derive a correction factor accounting for abiotic leaf mass loss and microbial decomposition. After 7 d of exposure, all G. fossarum, remaining leaf discs, and any visible leaf tissue were removed, dried at 60 °C to achieve a constant weight and subsequently weighed to the nearest 0.01 mg. The laboratory feeding trial with G. fossarum was performed to assess the ecotoxicological effects of the secondary (i.e. biologically) treated wastewater (referred to as “biology” in the remainder of the manuscript) released by WWTP Landau-Mörlheim while excluding confounding environmental factors that may have influenced the field experiments. In addition, the application of PAC (with and without nutrient amendment) as an advanced wastewater treatment technology was investigated for its potential to reduce adverse effects of organic as well as inorganic micropollutants (e.g. pharmaceuticals and personal care products; not quantified during the present study) in aquatic organisms. To achieve this, a 24-h wastewater composite sample was taken time-proportional (i.e. the same volume per unit of time) from the effluent of WWTP Landau-Mörlheim in February 2012 and stored

3 3 – – – 4 – – – 3 5 – – – – 1 – – – 1 – – – 5 – – – 1 9 – 33

LD3s (±2) (±1)

(±2)

(±1) (±5)

(±1)

(±0)

(±5)

(±0) (±4) (±7)

3 25 1 – – 5 – – – 6 7 – – – – 3 – 1 – – – – 1 – – – – 1 10 – 60

LD4s (±0) (±15) (±0)

(±3)

(±4) (±7)

(±2) (±0)

(±1)

(±0) (±5) (±30)

16 38 – – – 10 – – 10 50 92 17 – 1 – 6 – 8 – – – – – 1 – – – 5 3 – 258

LD5s (±5) (±22)

(±6)

(±5) (±16) (±58) (±17) (±0) (±3) (±6)

(±1)

(±1) (±3) (±73)

5 3 – – – 18 – – – 4 5 4 1 – – 2 – 1 – – – – – – – – – 6 38 – 86

(±3) (±2)

(±8)

(±2) (±4) (±3) (±0)

(±1) (±0)

(±5) (±20) (±10)

in stainless steel containers. Sixteen liters of sampled wastewater were treated in the laboratory with 20 mg PAC/L (contact time: 30 min). As PAC may remove ions essential for the leaf shredding amphipod (e.g. calcium; Bagrii et al., 2008) in addition to contaminants (i.e. (in)organic micropollutants), eight liters of PAC-treated wastewater (=PAC & nutrients) were subsequently amended with nutrients to an extent as present in the SAM-S5 medium (Borgmann, 1996) which is frequently used in amphipod laboratory assays (e.g. Zubrod et al., 2011). All wastewater samples were filtered (Whatman, GF/6, Beckenham, UK, pore size b 1 μm) to remove any particulate organic matter as well as PAC present and were then aerated for 24 h at 18 ± 1 °C until further use. For the laboratory feeding trial, one specimen of G. fossarum was placed together with two preconditioned leaf discs in a 250-ml-glass beaker filled with 200 ml of SAM-S5 (control), secondary- (biologically), PAC- or PAC & nutrients-treated wastewater. Beakers were aerated during the 7 days long experiment and placed randomized in a climate chamber at 20 ± 1 °C. For each treatment, 25 replicates were used with five additional beakers per treatment containing only two leaf discs to account for microbial decomposition and abiotic losses in leaf mass during the feeding trial. Amphipods, remaining leaf discs and any leaf tissue shredded off were removed after seven days of exposure, dried and weighed as described above. The feeding rate observed during the in situ bioassays as well as the laboratory feeding trial was calculated in milligram of consumed leaf mass per milligram dry weight of Gammarus per day (Maltby et al., 2000). 2.8. Calculations and statistics Shifts in the macroinvertebrate community composition, observed either by surber sampling or bycatch, were assessed by multivariate analyses for community ecology (vegan package for R; Oksanen et al., 2011). For these analyses, the species abundance data were not transformed to adequately display effects on common taxa (see Clarke and Warwick,

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correlation between environmental variables and sample similarities of the surber sampling data obtained during summer. Implications in leaf litter decomposition caused by wastewater released into the Queich were assessed separately for microorganisms and macroinvertebrates. The latter was calculated based on the leaf mass loss in coarse-mesh leaf litter bags corrected by the loss in fine-mesh leaf litter bags (type I). Macroinvertebrate as well as microorganism mediated leaf mass loss, fungal biomass, bacterial cell numbers and water parameters investigated during winter were assessed for significant differences between sampling sites by unpaired Student's t-tests or Wilcoxon rank sum tests. For summer data, ANOVAs followed by Dunnett's tests were applied. Abundance and biomass of gammarids in the bycatch during summer were checked for significant correlations with macroinvertebrate mediated leaf mass loss using Spearman's rank correlation. An ANOVA followed by Tukey's post-hoc test was performed to determine significant differences among treatments during the laboratory feeding trial. As every 7 day long in situ bioassay represents an independent experiment, the results were analyzed for significant differences among sampling sites using a fixed-effect meta-analysis (Borenstein et al., 2009), however, these analyses were conducted separately for the winter (n = 6) and summer (n = 3) seasons. The relative mean difference D (%) between the reference (=upstream) and the respective downstream site(s) was used as effect size. Additionally, all null hypothesis significance tests were supplemented by 95% confidence intervals (CI) (Altman et al., 2000) to obtain further information about effect sizes and precision. The proportions of dead gammarids in the in situ bioassays were compared among sampling sites using CI-testing with a Bonferroni adjustment for multiple comparisons. The term significant(ly) is exclusively used with reference to statistical significance (p b 0.05) throughout the present study. For statistics and figures, R version 2.13.2 was used (R Development Core Team, 2011). 3. Results 3.1. Wastewater dilution and water parameters

Fig. 2. Mean difference (with 95% CI; n = 5) in microorganism (a & b) and macroinvertebrate mediated (c & d) leaf litter decomposition between upstream (control) and downstream (below wastewater effluent) sites during winter (following 14 (□), 28 (○) and 42 d (Δ) of exposure) as well as summer (after 21 d). Asterisks denote significant differences, p b 0.05 (*), p b 0.01 (**), and p b 0.001 (***).

2001), especially gammarids, which are expected to contribute more to leaf litter decomposition (Dangles et al., 2004; Piscart et al., 2009, 2011) as compared to other more rare species. Non-metric multidimensional scaling (NMDS) ordination plots (Clarke, 1993) were constructed from resemblance matrices of Bray-Curtis similarities to provide a graphical interpretation of the community dissimilarities among sampling sites during both winter and summer respectively. Subsequently, surber sampling data were checked for homogeneity of group dispersions followed by a permutational multivariate analysis of variance (PERMANOVA) test to determine if there were significant differences in the community structure between sampling sites. A similarity percentage analysis (SIMPER; Clarke and Warwick, 2001) identified the species contributing most to dissimilarities in community structure. The envfit function, also included in the R package vegan (Oksanen et al., 2011), was used to assess the

At the inflow point, the wastewater contributed approximately 35 and 90% to the discharge during the winter and summer season, respectively. The upstream reference sites (LD1w, LD1s) differed with higher oxygen saturation (only during winter) and pH as well as lower values for temperature (only during winter), conductivity, ammonium, nitrate, nitrite, chloride and sulfate concentrations significantly from sites situated up to 300 m below the WWTP effluent. Differences in water quality parameters at sites located further downstream, i.e. 500 and 1000 m, were not significantly different compared to the reference sites (i.e. partly returning to reference conditions; Table A2). 3.2. Microbial parameters Fungal biomass associated with leaves deployed 100 m downstream of the WWTP effluent was significantly reduced by 43% when compared to the upstream reference site after 42 d of exposure during winter (Table A3). In the same sample, the number of bacterial cells was significantly increased by 137% (Table A4). During the summer, fungal biomass was significantly reduced at sites located 100 (by 76%) and 300 m (by 54%) below the WWTP effluent as compared to the reference site, while at sites 500 and 1000 m downstream, a non-significant increase (31 and 13%, respectively) was detected (Table A3). In contrast to the winter data, the number of leaf associated bacterial cells did not differ significantly among sampling sites (Table A4). 3.3. Microorganism and macroinvertebrate mediated leaf mass loss During winter, the microorganism mediated leaf litter decomposition was significantly increased by 5% (95% CI 2 to 8; p b 0.05; n = 5)

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and 8% (95% CI 3 to 13; p b 0.01; n = 4/5) at the downstream compared to the upstream reference site following 14 and 42 d of exposure (Fig. 2a). During the summer, in contrast, the downstream sites were not significantly different from the reference site (Fig. 2b). During the winter, macroinvertebrate mediated leaf litter decomposition at the downstream site (LD2w; located 100 m below the WWTP effluent) was significantly reduced compared to the upstream reference site by 43% and 65% following 28 and 48 d of exposure, respectively (Fig. 2c). During summer, the sites located up to 300 m downstream of the WWTP effluent consistently exhibited significant reductions in the amount of decomposed leaf litter (at 100 m by 28%; at 300 m by 33%; Fig. 2d). However, macroinvertebrate mediated leaf litter decomposition at sites 500 and 1000 m below the effluent was not significantly altered (Fig. 2d), i.e. comparable to the reference site. Differences in means between sites and corresponding CIs are provided in the Supplemental Data (Table A5). 3.4. Macroinvertebrate communities NMDS ordination showed that macroinvertebrate communities obtained by surber sampling were different between the upstream reference and the downstream site during winter (PERMANOVA: p b 0.05; n = 6; Fig. 3a). The differences in macroinvertebrate community composition, displayed by the NMDS ordination during summer, were more

pronounced between the reference site (but also 500 and 1000 m downstream) and sites located 100 and 300 m downstream the WWTP effluent (Fig. 3b). However, pairwise comparison of downstream sites with the reference site revealed no statistically significant differences (n = 3; PERMANOVA: p > 0.05). The envfit analysis showed that the closely correlated variables of nitrite, ammonium and temperature had a large influence (r2 > 0.99, p b 0.05, respectively) on the macroinvertebrate communities during summer. The SIMPER analysis revealed that the abundance of gammarids (especially G. fossarum and Gammarus pulex) at all downstream sites explained more than 35% of community dissimilarities compared to the respective reference site. At sites located 300 (LD3s) and 500 m (LD4s) downstream of the WWTP effluent, the increased abundance of the isopod Asellus aquaticus explained an additional 13 and 12% of the dissimilarities respectively (see Table 1 for further information). Although the SIMPER analysis revealed that Gammarus and Asellus accounted for 52 and 12% of the differences between LD4s and LD5s (n = 3; PERMANOVA: p > 0.05), respectively, (with higher abundances of both species at LD4s) these sites were more similar to each other than to LD2s or LD3s (Fig. 3b). Despite the lower level of taxonomic resolution, NMDS ordination and PERMANOVA, performed with the bycatch data (provided in Table A1), produced similar results for both seasons (PERMANOVA: p b 0.05; n = 4/5; Fig. 3c,d) compared to the analysis of the surber sampling data. Moreover, both the abundance (Spearman's rho = 0.88; p b 0.001)

Fig. 3. NMDS ordination plot of macroinvertebrate community data obtained by surber sampling (a & b; n = 6/3) and leaf bag bycatch (c & d; n = 5) during winter as well as summer. Similar symbols indicate replicates of a common sampling site.

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and biomass (Spearman's rho = 0.86; p b 0.001) of gammarids in the bycatch were significantly correlated with the macroinvertebrate mediated leaf litter decomposition during the summer. 3.5. In situ bioassays During the in situ bioassays performed in the winter season, the leaf feeding rate was significantly reduced 100 m downstream of the WWTP effluent discharge relative to the reference site (by 71%; Fig. 4). Similar observations were made during the summer season up to 500 m downstream of the effluent discharge (100 m: by 80%; 300 m: by 78%; 500 m: by 37%; Fig. 4). In addition, mortality amongst the gammarids was greater during the first week of the summer season at site LD2s (100 m below the WWTP effluent; difference of proportions = 75%; 95% CI 38 to 91%; p b 0.001; n = 20). Differences in means of the relative feeding rate between sites and corresponding CIs obtained by the meta-analysis are provided in the Supplemental Data (Table A6). 3.6. Laboratory feeding trial The feeding rate of gammarids exposed to secondary treated (=biology) wastewater in the laboratory was significantly reduced as compared to the control (by 59%; Fig. 5). Only the PAC treatment amended with nutrients improved gammarids feeding rate significantly (by 56%; Fig. 5) compared to the secondary treated wastewater. Means, corresponding CIs and p-values of differences between treatments, are provided in the Supplemental Data (Table A7). 4. Discussion Secondary treated wastewater released by the WWTP LandauMörlheim into the Queich reduced the leaf litter decomposition mediated by macroinvertebrates during both the winter and the summer season (Fig. 2c,d). Such implications in terms of this ecosystem function went along with pronounced alterations in the macroinvertebrate community composition at sites located 100 (LD2s) and 300 m (LD3s)

Fig. 4. Relative mean difference (with 95% CI) in gammarids' feeding rate obtained by a fixed-effect meta-analysis of weekly in situ bioassays data between the upstream and each downstream site during winter (Δ; n = 6) and summer (●; n = 3). Asterisks denote significant differences, p b 0.001 (***).

Fig. 5. Mean feeding rate (with 95% CI) of G. fossarum exposed to the control (=SAM-S5) and secondary treated (=biology) wastewater from the WWTP Landau-Mörlheim subjected at the lab-scale to PAC. PAC treated wastewater was additionally assessed following the amendment of nutrients (PAC & nutrients). Asterisks and horizontal lines denote significant differences between the respective two treatments, p b 0.05 (*).

downstream of the WWTP effluent, but given the reduced statistical power in the summer, only data for the winter season was statistically significant (Fig. 3a,b; cp. Schäfer et al., 2007). Moreover, the decrease in Gammarus (G. fossarum, G. pulex and G. roeselii) biomass and abundance positively correlated with the macroinvertebrate mediated leaf litter decomposition further emphasizing the key role of this genus in this ecosystem function (Dangles et al., 2004; Piscart et al., 2009, 2011). The observed dissimilarities in macroinvertebrate community composition may be explained by direct and indirect effect pathways (such as those relating to food quality) or a combination of multiple pathways caused by the release of the secondary treated wastewater. Although a significant increase in the abundance of leaf associated bacterial cells downstream of the WWTP was only observed during winter (LD2w; Table A4), fungal biomass was significantly reduced by more than 40% downstream of the WWTP effluent during both winter (100 m; LD2w) and summer (up to 300 m; LD2s/LD3s; Table A3). This could indicate potential shifts in the leaf-associated microbial community (Mille-Lindblom and Tranvik, 2003). Since fungi are particularly important for the palatability of leaf material for leaf shredding organisms (Chung and Suberkropp, 2009), food quality-related implications may partly explain the reduced abundance of this functional group (e.g. gammarids) in areas up to 300 m downstream of the WWTP effluent outfall. This hypothesis is also supported by a recent study by Zubrod et al. (2011) displaying that the consumption of leaf material exhibiting approximately 40% reduced fungal biomass can adversely affect the physiological fitness of leaf shredding organisms such as energy reserves and growth. This finally suggests that population development can be impaired if gammarids are forced to feed on leaf material with lower nutritious quality (Zubrod et al., 2011). However, as the in situ bioassays revealed significantly reduced gammarid feeding rates up to 500 m (LD2s to LD4s) downstream of the WWTP effluent (Fig. 4), direct adverse effects of the secondary treated wastewater are also conceivable. Since each of the in situ bioassays lasted only one week – an insufficient time span for meaningful deviations regarding microbial colonization (by organisms such as fungi) of the offered leaf discs (cp. Hieber and Gessner, 2002) – these bioassays can be assumed to display most likely direct, not food quality-related, implications

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(Bundschuh and Schulz, 2011) of wastewater exposure. Moreover, the decrease in feeding rate of up to 80% during winter as well as summer season (Fig. 4) suggests a reduction in energy reserves and a reduced growth of individuals (Bundschuh et al., 2011c), which would affect population development at impacted sites (Baird et al., 2007). Most importantly, the in situ bioassays detected a significantly reduced feeding rate 500 m (LD4s) downstream of the WWTP effluent within the Queich's main stream, while the macroinvertebrate-mediated leaf litter decomposition in coarse-mesh leaf litter bags was not different from the upstream site (Fig. 2c,d). This discrepancy may be explained by the relatively high abundance of shredders such as Asellus sp. and Gammarus sp. (Table 1), at this site (LD4s) potentially compensating for their impaired feeding rate on leaf material as indicated by the in situ bioassays (Fig. 4). Although agricultural land-use may have also affected the implications reported at site LD4s, this scenario is – due to the broad buffer strips – very unlikely. The markedly altered levels of water quality parameters at impacted sites (Table A2), comparable with the undiluted effluent sample (Table A8), further support the hypothesis of direct effects. Stream temperature, nitrite and ammonium levels correlated significantly with alterations in macroinvertebrate community composition. However, the nutrients ammonium and nitrate are reported to be relatively non-toxic for the dominant shredder genus (i.e. Gammarus) within the investigated stream section (Stelzer and Joachim, 2010; Williams et al., 1986). Moreover, growth rates of offspring, juvenile and adult gammarids increase with rising water temperatures up to 20 °C (Pöckl, 1992; Sutcliffe et al., 1981), which was not exceeded during the present study (Table A2). The levels of unionized ammonia measured in the present study also remained nearly 50-fold below the chronic no observed effect concentration for G. pulex (Berenzen et al., 2001). In addition, the nitrite concentrations measured in the present study were a factor of 5 to 10 below those acutely toxic for other amphipods, i.e. Eulimnogammarus toletanus or Hyalella atzteca (Alonso and Camargo, 2009; Soucek and Dickinson, 2012). Nevertheless, chronic effects of the measured nitrite concentration cannot be excluded (see Berenzen et al., 2001). The sublethal effects of nitrite – potentially in combination with the other nutrients released – may have driven the low abundance of gammarids below the WWTP effluent and were possibly aggravated by the low oxygen levels (Table A2). However, the impact of the latter may be limited as the laboratory feeding trial, which was performed under permanent aeration, revealed an impairment in the feeding rate of G. fossarum exposed to secondary treated wastewater at a comparable effect size (approximately 60%; Fig. 5; Table A8) as observed during the in situ bioassays (Fig. 4). In addition to nutrients and potentially oxygen, several (in)organic micropollutants (not quantified during the present study), which are commonly discharged by WWTPs (Bueno et al., 2012; Daughton and Ternes, 1999; Dyer and Wang, 2002), may have directly affected the macroinvertebrate community composition. For instance, adverse implications of secondary treated municipal wastewater from a WWTP in Switzerland were recently linked to the presence of micropollutants (Bundschuh and Schulz, 2011), supporting the importance of these substances in the outcomes seen in the present study. Considering the decreased dilution potential of receiving streams (particularly during summer droughts) presently and with the predicted accentuation under global climate change scenarios (Prellberg, 2010), the adverse implications of secondary treated wastewaters containing multiple micropollutants may become even more pronounced. Although the dilution potential of the Queich branch was 2.6-fold higher during the experiments performed in the winter compared to the summer season neither the decomposed leaf litter nor the in situ measured feeding rate of G. fossarum detected deviations in the effect sizes among seasons at the site located 100 m (LD2w/s) downstream of the WWTP effluent. However, as the winter experiments are lacking multiple sites, no information about the distance impacted by the wastewater released during scenarios of differing dilution potential can be derived. Nevertheless, it may be assumed that the area of a stream section negatively affected by a WWTP effluent is reduced

at times of higher water level such as in winter, and consequently a higher dilution of introduced wastewater (cp. Birge et al., 1989). Since the present study is limited to a single stream, future studies should consider multiple streams allowing for more general conclusions on the implications of conventionally treated wastewater in freshwater ecosystems during low flow conditions and hence climate change scenarios. However, to counteract such adverse effects of insufficiently diluted wastewater containing (in)organic micropollutants in aquatic ecosystems, advanced wastewater treatment technologies may be applied (cp. Bundschuh et al., 2011a). In this study, the potential of PAC to reduce the toxicity of the secondary treated wastewater by adsorbing micropollutants was investigated. The laboratory feeding trial with G. fossarum suggests that the application of PAC resulted in reduced toxicity of the secondary treated wastewater if nutrients (according to the SAM-S5 medium; Borgmann, 1996) were amended following the PAC treatment (Fig. 5). The absence of an ecotoxicologically beneficial effect without the amendment of nutrients may be explained by the capability of PAC to adsorb nutrients like calcium ions (Bagrii et al., 2008), which are essential for development of crustaceans (Neufeld and Cameron, 1993). The loss of these calcium ions hence may have overridden the positive implications resulting from the adsorption of the broad range of (in)organic micropollutants usually present in wastewater by the PAC treatment (Bundschuh et al., 2011d). However, as such essential ions are available in abundance in the receiving stream, nutrient amendment – as conducted during the present study – would not be necessary in the field. Hence, the application of PAC as an advanced wastewater treatment step seems to be a promising tool to reduce the ecotoxicological potential of conventionally treated wastewater (cp. Bundschuh et al., 2011d).

5. Conclusions In conclusion, wastewater released from a conventional municipal WWTP affected microbial conditioning of leaf material, macroinvertebrate feeding as well as abundance, and fundamental ecosystem functions (leaf litter decomposition) regardless of the seasonal differences in stream water levels of the investigated stream. These implications suggest a reduction in the energy supply in terms of fine particulate organic matter for local and downstream communities (e.g. collectors), potentially also affecting representatives of higher trophic levels such as predators (Wallace et al., 1997). Considering the increased aggravation of water scarcity, in conjunction with a reduced wastewater dilution potential of streams as predicted by global climate change scenarios (Prellberg, 2010), our results may in the future also be of relevance for regions in Europe, which are currently still water-rich. Advanced wastewater treatment technologies as applied in a worst case scenario (i.e. no dilution of wastewater) during the present study may help to protect the integrity of freshwater ecosystems and to achieve appropriate water quality standards as required by the European Water Framework Directive (European Commission, 2000).

Formal assurance The authors assure that the present study was conducted in accordance with national and institutional guidelines for the protection of animal welfare.

Conflict of interest We declare that there is no conflict of interest with any financial organisation regarding the material discussed in the manuscript.

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Acknowledgments The authors thank P. Baudy, R. Bundschuh, T. Bürgi, K. Englert, A. Feckler, M. Konschak, R. R. Rosenfeldt, R. S. Schulz, F. Seitz and A. Stoepel, for help in the laboratory and field, E. Szöcs for advice in multivariate statistics as well as E. Brockmeier for proof reading of this manuscript. Moreover, the Fix-Stiftung, Landau is acknowledged for financial support of research infrastructure.

Appendix A. Supplementary data Supplementary data to this article can be found online at http:// dx.doi.org/10.1016/j.scitotenv.2013.03.025.

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