leaf decay processes during and after a supra-seasonal hydrological drought in a temperate lowland...

© 2011 WILEY-VCH Verlag GmbH & Co. KGaA, Weinheim 1434-2944/11/611-0633

Internat. Rev. Hydrobiol. 96 2011 6 633–655

DOI: 10.1002/iroh.201111322

JEANETTE SCHLIEF* and MICHAEL MUTZ

BTU Cottbus, Department of Freshwater Conservation, Seestraße 45, 15526 Bad Saarow, Germany; e-mail: [email protected]

Research Paper

Leaf Decay Processes during and after a Supra-Seasonal Hydrological Drought in a Temperate Lowland Stream

key words: hydrological drought, climate change, litter decomposition, microbial activity, shredder

Abstract

Climate change models for Central Europe predict hydrological drought with fragmentation into pools during periods of high litter input in numerous lowland streams, presumably affecting in-stream leaf decay processes. To investigate this assumption, we measured physicochemical parameters, macro-invertebrate colonization, microbial activity, and decay rates of exposed leaves during and after a supra-seasonal drought in a German lowland stream. Microbial activity, shredder colonization and leaf decay rates during fragmentation were low, presumably caused by drought-related environmental conditions. Microbial activity and temperature-corrected decay rates increased after the flow resumption but not leaf mass loss and shredder colonization. During both periods, exposed leaves appeared physically unaf-fected suggesting strongly reduced shredder-mediated leaf decay despite shredder presence. Our results indicate that hydrological drought can affect organisms and processes in temperate lowland streams even after flow resumption, and should be considered in climate change scenarios.

1. Introduction

The decomposition of allochthonous organic matter, such as leaf litter from riparian trees, is a major ecosystem-level process in streams running through forested watersheds (WEB-STER and BENFIELD, 1986; ABELHO, 2001). Leaves are processed by a complex interaction of several abiotic and biotic processes, such as physical leaching, mechanical abrasion, microbial degradation, and consumption by macro-invertebrates (leaf shredders). Aquatic hyphomycetes play a predominant role in microbial leaf decay (HIEBER and GESSNER, 2002; PASCOAL and CÁSSIO, 2004), as they rapidly colonize submerged leaves, degrade plant cell polymers (CHAMIER, 1985) and increase the palatability of leaves for shredders (GRAÇA and CANHOTO, 2006). Fine organic particles and dissolved compounds resulting from leaf decay serve as energy sources for various consumers (BENFIELD, 1996). Hence, understanding the responses of in-stream leaf decay processes to climate change is central to interpreting and predicting the impacts of climate change on stream ecosystems.

Higher temperatures, locally reduced precipitation and increased frequency and duration of extreme events (e.g., seasonal and supra-seasonal hydrological droughts) are expect-

* Corresponding author

634 J. SCHLIEF and M. MUTZ

© 2011 WILEY-VCH Verlag GmbH & Co. KGaA, Weinheim www.revhydro.com

ed consequences of climate change in temperate climate regions, such as Central Europe (IPCC, 2007). These conditions may cause regional losses in the total hydrological budget (GERSTENGARBE et al., 2003) with implications for the annual discharge regime of rivers and streams (KRYSANOVA et al., 2008), resulting in the temporary drying of first- to third-order lowland streams, which currently have permanent flow, throughout the summer and possi-bly extending throughout the autumn (ANDERSEN et al., 2006). Such drying can disrupt the hydrological connectivity of a stream (LAKE, 2003) leading to watercourse fragmentation or even complete streambed drying (BOULTON, 2003). In lowland streams, which are mostly groundwater-fed and characterized by sandy substrates, droughts may fragment the stream into a series of isolated groundwater-sustained pools.

Riparian vegetation stress during drought in temperate regions with deciduous forests can cause premature leaf abscission (WENDLER and MILLARD, 1996; KOZLOWSKI and PALLARDY, 2002), resulting in earlier litter input into streams. During summer drought, relatively large amounts of leaf litter could, therefore, be dispersed in fragmented streambeds during a sea-son commonly characterized by low litter supply. If the hydrological drought is extreme and lasts throughout the autumn, the total autumnal leaf litter input will accumulate in a dried up or fragmented streambed and will be supplied to the total stream communities when flow resumes later in the year with unknown effects on leaf decay processes.

In temperate climate regions characterized by an autumn pulse of leaf fall, litter is com-monly supplied and decomposed during periods of high discharge and low temperature of streams. Shifts in the timing, quality and amount of the litter supply may have great effects on leaf-associated organisms, including leaf shredders and aquatic fungi well-adapted to performing leaf decay under these autumnal conditions. Shredders are known to syn-chronize their life cycles to maximize the use of the seasonally varying leaf litter supply (CUMMINS et al., 1989). Aquatic hyphomycetes show their best development in response to autumnal leaf fall (CHAUVET, 1991; BÄRLOCHER, 1992) and have maximum sporulation and biomass in autumn and winter (FABRE, 1998; METHVIN and SUBERKROPP, 2003).

Studies on litter decomposition and associated organisms in intermittent streams are scarce and are mainly performed in warmer and drier climate regions (MAAMRI et al., 1998; LAKE, 2003; ACUÑA et al., 2005). In these streams, fragmentation is accompanied by abrupt changes in physicochemical conditions (e.g., increased water temperatures and decreased oxygen lev-els), which have adverse effects on stream organisms (GASITH and RESH, 1999; CANHOTO and LARANJEIRA, 2007). Moreover, drought inhibits detritus breakdown (LAKE, 2003), changes spatial processing patterns (PINNA et al., 2004) and alters the importance of macro-inverte-brates in litter processing (MAAMRI et al., 2001). The transferability of these findings to other climatic zones is uncertain. Studies considering the effects of extreme drought events on leaf decay processes in temperate lowland streams, which are rarely subject to drying, are almost completely lacking. Increased litter input through premature or peak leaf fall and increased retention time during drought may increase the availability and palatability of detritus to detritivores. At the same time, reduced stream area and depth during droughts will decrease the habitat space and increase the vulnerability of leaf-decaying organisms through compe-tition and predation, which could cause leaf decay rates to decrease (SMITH and PEARSON, 1987; SMOCK et al., 1994). Indirect effects of droughts, such as deteriorated water quality, could also affect stream organisms and, in turn, leaf decay rates. Moreover, knowledge about leaf decay processes after flow resumption of temperate lowland streams is almost lacking. The few studies assessing post-drought periods in temperate streams observed partly high macro-invertebrate recolonization rates, but also unsuccessful recruitment of specific taxa (BOULTON, 2003). However, to what extent recolonization dynamics affect in-stream leaf decay processes after an extreme drought event in temperate lowland streams is unknown.

Our objective was to assess the response of leaf decay processes to an extreme hydrologi-cal drought event in a temperate lowland stream within two periods, during fragmentation and directly after flow resumption. We assumed that 1) during fragmentation, drought related

Leaf Decay during Drought 635


environmental conditions affect leaf associated organisms and, in turn, decay processes, and 2) after flow resumption, when environmental conditions have improved, but leaf-associated biocoenosis is eventually still impoverished, leaf decay processes may still be affected. Phy-sicochemical parameters, leaf associated macro-invertebrate colonization, microbial activity, and leaf decay rates in a lowland stream were compared during stream fragmentation and after flow resumption to provide a context for understanding and predicting the drought responses of leaf decay processes.

2. Methods

2.1. Study Site

Extremely low regional rainfall combined with high air temperatures between June and October 2006 in Central Europe lowered the water tables and created extensive areas of groundwater recharge. During this severe meteorological drought within a region usually characterized by permanent streams the second-order stream “Verlorenwasser” provided ideal conditions to perform a case study of drought effects in a temperate lowland stream. The Verlorenwasser is located in Brandenburg, Germany, 60 km southeast of Berlin, within the catchment area of the River Spree. The Verlorenwasser is an unregu-lated, low gradient, sand-bed stream running through a deciduous forest with alder (Alnus glutinosa [L.] GAERTN.) and oak (Quercus robur L.) as the dominant riparian tree species. Along its course, the Verlorenwasser flows through a wetland (area: ~1 ha, mean water depth: ~1 m). The stream flow of the Verlorenwasser is either permanent with summer/autumn low flow periods, or intermittent with short (~ 4 to 6 weeks) seasonal hydrological drought with fragmentation into pools of large stream sections in summer or early autumn depending on annual or seasonal precipitation (SCHLIEF and MUTZ, unpublished data). Three sampling sites that are known to develop into remnant pools during fragmentation (SCHLIEF, pers. observ.) were selected within a 500 m long reach of Verlorenwasser (sites: S 1, S 2 and S 3). This reach is located downstream of the wetland and has a slope of 0.2%. The pools during fragmentation were similar in size (~3–3.5 m2) but differed in water depth.

2.2. Physicochemical Parameters

Weekly to biweekly inspections of the study reach were performed to observe alterations in the flow regime of the Verlorenwasser (e.g., onset of stream fragmentation or flow resumption) between April 2006 and April 2007. Data loggers (PL-01, GERO Messsysteme GmbH, Germany) continuously recorded water level, and water and air temperature at all sites during this period. The water level recorded was inspected for extreme drought conditions indicated by water level decreases (< 2 cm above the streambed level). Flow conditions in the whole stream and the proportion of fragmentation in rela-tion to the total stream length were assessed at monthly to bi-monthly intervals. All other parameters were measured weekly during two five-week periods of leaf pack exposure. Conductivity and pH were measured using portable meters (WTW LF 323, WTW pH 325). Dissolved oxygen (DO) concentrations in the water column, in the surface layer and in the centre of the leaf packs were measured using fibre optic oxygen microsensors (PreSens GmbH). Water samples (minimum 1 litre) for NO3–N, NH4–N, soluble reactive phosphorus (SRP), total phosphorous (TP), and total nitrogen (TN) analyses were randomly taken using a 1.5 litre plastic jar. Samples for nitrate, ammonium and SRP analyses were fil-tered through cellulose-acetate membrane filters (SARTORIUS, 0.45 μm pore size), and determined by segmented flow analysis (PERSTORP-Analytical). The TP and TN concentrations in unfiltered samples were determined by flow analysis (FIA). All water samples were stored in sealed plastic jars at –20 °C until analysis. Current velocity during periods of permanent stream flow was measured using a small (1.5 cm diameter) hydrometric propeller (Mini-Air-2, Schiltknecht), which was positioned in ~50% of the water depth. Current velocity during fragmentation was measured by acoustic Doppler velocimetry (ADV), as described by MUTZ (2000). Groundwater inflow at sampling sites was determined by using 30 small plastic piezometers installed randomly with one end inserted in the stream bed and the other end extending above the water level, and then the water surface height in the piezometer was measured relative to that of the surrounding water.



2.3. Leaf Pack Preparation, Exposure and Retrieval

In November 2005, freshly fallen alder leaves were collected during leaf abscission using nets hang-ing over the Verlorenwasser stream channel and were stored dry at room temperature until needed. Alder leaves were expected to provide considerable litter input during a potential fragmentation period in the Verlorenwasser, since it is a dominant riparian tree species with typically large quantities of leaf fall in summer and early autumn (KIRUZAWA, 1980). In July 2006 and February 2007, leaves were aliquoted into 5 g packs (consisting of 9.8 ± 1.9 alder leaves) and placed in coarse (mesh size: 20 mm, area: 20 × 20 cm, height: 2 cm) bags 24 h prior to exposure. Holes 40 mm in diameter were made in the top of the bags to allow the entry of larger macro-invertebrates. Once stream fragmentation and watercourse disruption had occurred, the first leaf exposure experiment was started on 5 July 2006. Thirty leaf packs were placed in each site (90 packs total), corresponding to ~50 g leaf litter per m2 stream bottom and simulating premature high litter input as would be expected to occur due to climate change. The leaf packs were exposed directly at the natural litter layer and secured to the streambed using tent pegs and nylon cord. Three additional leaf packs were not submerged in the stream but were returned to the laboratory to determine initial mass.

After 7, 14, 21, 28, and 35 days of exposure, six randomly selected packs per site (18 packs total) were carefully retrieved using a hand net (250 μm mesh size), placed in polyethylene boxes half-filled with stream water, and transported in a cooler to the lab. After flow had resumed in the stream, a second leaf exposure experiment was started on 26 February 2007 under identical test conditions. Accordingly, there were two experimental periods: period 1 (during fragmentation) and period 2 (after flow resumption).

2.4. Leaf Mass Loss

The remaining leaf material of all sampled packs was carefully inspected in the lab for shredder feeding signs (e.g., skeletonization or holes at the margin or lamina of leaves) and rinsed with filtered stream water (10 μm filter) over a series of sieves (1 mm, 500 and 250 μm) to remove particles and macro-invertebrates. Leaf material retained by the 1 mm sieve from three randomly selected packs per site was then used for mass loss determination and weighed after drying (48 h at 60 °C) and ashing (4 h at 550 °C) to determine the ash-free dry mass (AFDM). Leaf mass loss data were expressed as a percentage remaining from the initial ash-free dry mass, which was determined according to the same protocol.

2.5. Macro-Invertebrates

All macro-invertebrates detected during the rinsing of the leaf material or retained by the sieves were preserved in 70% ethanol and later counted, classified to the lowest practical taxonomic level and divided into two groups – shredders or non-shredders – according to TACHET et al. (1987). Invertebrates were dried (48 h at 60 °C), weighed, ashed (550 °C, 4 h), and reweighed for AFDM determination. The abundance and biomass of shredders and non-shredders per leaf pack were expressed per unit weight of leaf AFDM.

2.6. Leaf-Associated Microbial Activity

The rates of leaf associated respiration and fungal sporulation as measures of microbial activity were assessed within five to six hours after sampling. For this purpose, the leaf material of the three packs per site that were not used for mass loss determination was divided into two portions: One portion (~ five leaves per pack) was used to measure leaf-associated respiration based on oxygen consumption, and one portion (~ two leaves per pack) was used to induce fungal sporulation. Whole leaves were used for the determination of microbial activity to comprise the total microbial community associated with all leaf components (including petioles, leaf veins and margins).

For the respiration measurement, leaves were placed in autoclaved 100 ml glass bottles (three repli-cates per site), each with a sintered neck, bevelled stopper and small glass marble inside. Additionally,



an inflexible plastic mesh was placed inside the bottles to prevent contact between the leaves and the marble. Subsequently, the bottles were filled with filtered (10 μm) stream water and closed without air bubbles inside. The bottles were placed on an orbital shaker (100 rpm) inside a climate chamber. The gla ss marbles ensured continual water movement inside the bottles. The oxygen consumption inside the bottles was then measured continuously using oxygen microsensors affixed inside the bottles and connected via plastic glass-fibre cable to an oxygen optode (Microx T, Precision Sensing GmbH). The climate chamber had a constant temperature and was dark to prevent photosynthesis. The target was to measure actual rather than potential leaf-associated respiration rates. Thus, temperatures and initial oxygen levels (by bubbling with air or nitrogen) were adjusted to approximate the present in-stream conditions of the corresponding sampling date, and ranged from 17 to 21 °C and 2.8 to 3.5 mg l–1 oxy-gen during fragmentation, and from 6.0 to 6.5 °C and 9.5 to 12.5 mg l–1 oxygen after flow resumption. Blanks with filtered (10 μm) stream water only were measured concurrently and used for the correction of leaf-associated respiration rates. It took between one and two hours to complete the measurements depending on the level of activity. Afterwards, the leaves were carefully retrieved, dried and ashed according to the same protocol used for mass loss determination. The respiration rates were expressed as mg oxygen consumed per unit ash-free dry weight of leaves per hour.

To induce fungal sporulation, leaves were placed in 250 ml Erlenmeyer flasks (three replicates per site) containing 100 ml of filtered stream water (Whatman GF/D filter, 0.45 μm pore size) and incubated on an orbital shaker (100 rpm) for two days at 15 °C. After incubation, 250 μl of 0.5% (w/v) Triton X-100 solution was added and 50 ml of the conidial suspension was passed through a membrane filter (5 μm pore size). Conidia trapped on the filter were stained with 0.1% cotton blue in 60% lactic acid, identified (INGOLD, 1975; WEBSTER and DESCALS, 1981; GULIS et al., 2005) and counted at a magni-fication of 200X to 400X. Two filters per sample were prepared and at least 25 fields per filter were counted. Additionally, whole filters were scanned for rare taxa. After sporulation, the leaves were care-fully retrieved, dried and ashed according to the same protocol used for mass loss determination. The sporulation rates were expressed as the number of conidia released per unit leaf ash-free dry weight per day.

2.7. Data Analysis

Physicochemical water parameters, macro-invertebrate abundance and biomass, respiration rates, sporulation rates, and leaf mass remaining were used as dependent variables and compared between periods (period 1 vs. period 2) with ANCOVA (GLM procedure), with period as factor and site, and exposure time and exposure to extreme drought conditions as covariates. If significant influences of covariates were detected, we performed an ANOVA for each period separately to test in which period effects of site, exposure time or extreme drought conditions on the dependent variables were present. The leaf decay rates were calculated from linear regressions on ln-transformed mass loss data based on either days of exposure (k) or degree-days (k') to account for temperature differences (BOULTON and BOON, 1991), and ANCOVA (GLM procedure) was used to test for differences in decay rates between periods. Data were log-transformed when necessary to normalize the data and homogenize the variances. Means (± standard error) presented in text and figures were calculated using non-transformed data. All statistical analyses were performed using SPSS (Vol. 14.0/SPSS Inc., Chicago, IL) with significance levels set at P < 0.05.

3. Results


From late-June 2006 until mid-January 2007, hydrological drought caused fragmentation with pool formation in the upper and middle reaches of the Verlorenwasser (~75% of total stream length) and complete drying of the lower reach (~25% of total stream length) with-out pool formation. Even the wetland located upstream of the study reach dried up leaving behind an isolated water body (area: ~50 m2). As the hydrological drought lasted more than six months, it can be classified as a supra-seasonal drought. The water level in the



stream continued to decrease during the initial two weeks of leaf pack exposure in period 1 (Fig. 1a), leading to extremely low average pool depths of less than 30 cm at S 1 and S 2, and less than 12 cm at S 3. The pools were initially connected by a small channel (width 5 to 15 cm, water depth ~2–4 cm), which disappeared during the second week, leaving behind a series of isolated pools. There was a noticeable discrepancy in the mean water temperature of week one (20.2 °C) and weeks three to five (17.4 °C), coinciding with the periods before and after the pool isolation (Fig. 1c). We also observed daily water level fluctuations of 2 to 10 cm per day in the pools during fragmentation (Fig. 1a), with a maximum water level between midnight and early morning periods. The occurrence of extreme drought conditions were detected in one pool (S 3), since after the first week, the daily fluctuations temporarily lowered the water level at this site to less than 2 cm above the streambed level during after-noon and early night periods for three to ten hours per day. The minimum water level was achieved for one to two hours during late afternoon (~16:00 to 18:00 pm) and ranged from 1 cm (weeks two – four) up to 2 cm (week five) below the streambed level at S 3. Visual inspection during periods of minimum water level at S3 revealed that the natural litter layer and the exposed leaf packs remained waterlogged and did not become dry, but the uppermost leaves were only covered by a thin water film. Most of the water parameters at this site were measured when the water level was ~5 to 15 cm above streambed level, and oxygen profiles at S 3 could only be measured during the initial three weeks. The deeper pools (S 1 and 2) continuously contained water with a water level > 10 cm above the streambed level during fragmentation (Fig. 1a). Water depths at the sites after flow resumption were up to four times higher than those during fragmentation, and no daily fluctuations were measured. The water level after flow resumption was not measured continuously at S 1 due to a techni-cal problem, but a similar overall pattern to that at S 2 and S 3 can be assumed (Fig. 1b). At the beginning of the exposure period during fragmentation, natural leaf accumulations at the stream bed were sparse and consisted mainly of a very few brown oak leaves, presum-ably shed in the previous autumn. As the study proceeded, newly shed green alder leaves increasingly spread over the stream bottom. After flow resumption, the whole stream bottom was covered by thick layers of leaf litter (mainly oak and, to a lesser degree, alder leaves).

The mean flow velocity in the pools was extremely low during fragmentation, namely, 0.9 to 1.4 cm s–1 initially, and 0.6 to 0.7 cm s–1 after two weeks. After flow resumption, there was permanent stream flow at all sites, with higher (ANCOVA, P < 0.05) mean flow velocities (10 to 15 cm s–1) than during fragmentation, as expected. During fragmenta-tion, extremely low DO concentrations within the hypoxic range were measured in the pools without appreciable differences between the water column, centre and surface layer of the leaf packs (Table 1). Although water in the pools was not stagnant, oxygen profiles revealed a gradient of highest values near the water surface and lowest close to the stre-ambed (Fig. 2a), indicating groundwater influence in the pools during fragmentation. In addition, the water level in all piezometers was higher (0.2 to 0.7 cm) than the surrounding stream water level, indicating groundwater inflow. After flow resumption, oxygen concentra-tions at all sites were close to saturation without a depth gradient (Fig. 2b) and significantly higher (ANCOVA, P < 0.001) in the water column, the surface layer and centre of the leaf packs than the corresponding concentrations during fragmentation. As the exposure period after flow resumption started in winter, the mean daily water temperature was 11 to 12 °C lower (ANCOVA, P < 0.001) than during fragmentation (Table 1). SRP, NH4–N, TP, and conductivity were lower (ANCOVA, P < 0.05), whereas NO3–N was up to 76 times and TN up to 9 times higher (ANCOVA, P < 0.001) after flow resumption than during fragmentation (Table 1). The pH was circumneutral in both periods.



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Figure 1. Water level (a) and water and air temperatures (b) measured at the Verlorenwasser during two five-week leaf pack exposure periods. Data shown are (a) continuous water level measurements at the three sampling sites (S 1–3) and (b) means of constant temperature measurements at site 2 during

either period 1 or 2.

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3.2. Macro-Invertebrates

ANCOVA revealed no effect of the exposure period either on the abundances of shredders or non-shredders (ANCOVA, P = 0.83 and P = 0.32, respectively), or on the corresponding biomasses (ANCOVA, P = 0.30 and P = 0.53, respectively). Four macro-invertebrate spe-cies found were classified as shredders: the amphipod Gammarus pulex, the isopod Asellus aquaticus and two Trichoptera species, Allogamus uncatus and Glyphothaelius pellucidus (Table 2). A. aquaticus was the dominant leaf shredder in both periods. Only a few indi-viduals of G. pulex were associated with the leaf packs on the first sampling date during fragmentation (Table 2). At the same time, several dead G. pulex individuals were observed at the stream margins. No G. pulex individuals, but a few Trichoptera individuals were found in the leaf packs after flow resumption (Table. 2). During fragmentation, shredder abundance decreased after the first week and remained at a low level with a mean abundance of only one to four individuals per g AFDM (Fig. 3a). We observed a similar pattern for the leaf-associated biomass of shredders (Fig. 3c). After flow resumption, shredder abundance and biomass remained at an extremely low level with an average of zero to five individuals per g AFDM during the entire exposure period.

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Figure 2. Depth profiles of oxygen concentration during two five-week leaf pack exposure periods in the Verlorenwasser. Data shown are means ± standard errors of three measurements during period 1 (a)

and five measurements during period 2 (b) at the sampling sites.



Overall, there was no effect of sampling site or extreme drought conditions on the abun-dances and biomasses of shredders and non-shredders, although on one occasion (week four) at one site (S 3) during fragmentation, shredder abundance and biomass increased again up to three to four times higher (Fig. 3a/c). This might have been caused by numerous small (< 3 mm) A. aquaticus individuals, which appeared only on this sampling date. Moreover, on one occasion (week one) at S 1 after flow resumption, the presence of large Trichoptera individuals caused a slightly higher biomass.

Predatory taxa were identified among the non-shredders, e.g., Platambus maculatus, Nepa cinerea and three taxa of Hirudinea (Table 2), during both periods. The abundance and bio-mass of non-shredder taxa followed similar patterns at all sites, with comparably low values (average of zero to four individuals per g AFDM) during both exposure periods (Fig. 3b/d).

3.3. Leaf-Associated Microbial Activity

ANCOVA revealed a significant effect of the exposure period on leaf-associated respira-tion (ANCOVA, P < 0.001) and sporulation (ANCOVA, P < 0.05). Leaf-associated respira-tion rates (Fig. 4a) were lower during fragmentation than after flow resumption. Further-more, the temporal pattern of respiration rates differed between the two exposure periods. At all sites during fragmentation, respiration rates decreased from an initial maximum of ~0.8 mg O2 h–1 g–1 AFDM to half that amount, and then slightly increased towards the end of the exposure period. At all sites after flow resumption, respiration rates showed an inverse pattern characterized by an initial increase to a maximum of ~2 mg O2 h–1 g–1 AFDM at week three and a later decrease (Fig. 4a).

Similar to respiration, sporulation rates were lower during fragmentation than after flow resumption (Fig. 4b). At all sites during fragmentation, almost no spores were counted until the last sampling date, when only four to five taxa (Table 3) releasing around 0.04–0.1 conidia μg–1 leaf AFDM d–1 were detected (Fig. 4b). The highest numbers of conidia during fragmentation were released by Culicidospora gravida, Flagellospora curvula and Helis-cella stellata. All spore-releasing taxa detected during fragmentation were also found after

Table 1. Physicochemical characteristics at three sites (S 1─3) in the Verlorenwasser dur-ing leaf pack exposure in period 1 (fragmentation) and period 2 (flow resumption). Values

represent means ± standard error.

Period 1 Period 2

S 1 S 2 S 3 S 1 S 2 S 3

Temp. water (°C) 17.0 ± 0.4 17.4 ± 1.9 18.2 ± 2.7 6.0 ± 1.2 6.0 ± 1.1 5.9 ± 1.1Temp. leaf pack surface (°C) 17.8 ± 2.4 17.9 ± 2.6 17.7 ± 1.4 6.3 ± 0.2 6.3 ± 0.3 6.4 ± 0.3Temp. leaf pack center (°C) 17.3 ± 2.3 17.5 ± 0.9 17.5 ± 1.7 6.3 ± 0.2 6.3 ± 0.2 6.3 ± 0.2O2 water (mg l–1) 1.7 ± 0.4 1.6 ± 1.5 1.9 ± 1.2 10.2 ± 0.7 10.1 ± 0.8 10.3 ± 0.7O2 leaf pack surface (mg l–1) 1.5 ± 1.0 1.4 ± 0.9 1.8 ± 1.1 10.1 ± 0.5 10.2 ± 0.5 10.2 ± 0.7O2 leaf pack center (mg l–1) 1.0 ± 0.2 0.9 ± 0.6 1.1 ± 0.4 10.1 ± 0.7 10.1 ± 0.8 10.2 ± 0.6Flow velocity (cm s–1) 0.70 ± 0.04 0.71 ± 0.04 0.67 ± 0.03 12.7 ± 4.1 10.4 ± 3.7 15.1 ± 4.9pH 7.2 ± 0.5 7.0 ± 0.5 7.2 ± 0.3 7.5 ± 0.4 7.5 ± 0.3 7.5 ± 0.3Conductivity (μS cm–1) 963 ± 112 988 ± 106 981 ± 111 749 ± 42 738 ± 30 735 ± 34SRP (μg l–1) 235 ± 90 270 ± 108 313 ± 280 7.4 ± 1.6 7.3 ± 1.9 7.4 ± 1.7NO3–N (mg l–1) 0.2 ± 0.2 0.2 ± 0.2 0.2 ± 0.2 12.7 ± 2.6 12.7 ± 2.2 12.6 ± 2.4NH4–N (μg l–1) 948 ± 278 600 ± 288 683 ± 248 69 ± 10.0 67.2 ± 9.8 69.9 ± 10.3TP (μg l–1) 314 ± 68 355 ± 30 450 ± 137 45.8 ± 4.9 42.5 ± 5.4 44.9 ± 5.9TN (mg l–1) 2.0 ± 0.8 1.6 ± 0.5 1.8 ± 0.2 13.9 ± 3.2 13.9 ± 2.8 13.3 ± 3.8



flow resumption, with the exception of Flagellospora curvula (Table 3). Sporulation rates after flow resumption showed a temporal pattern similar to that of respiration rates within the same period (Fig. 4b): An initial increase to a maximum of around 2.2–2.5 conidia μg–1 leaf AFDM d–1 released by up to nine taxa was observed at weeks three and four at all sites, followed by a subsequent decrease. A total of up to 13 to 16 spore-releasing taxa were detected after flow resumption (Table 3). Alatospora acuminata, a species not detected during fragmentation, contributed more than 60% of the total conidia released after flow resumption, and Culicidospora gravida and Heliscella stellata also released high numbers of spores (Table 3).

Table 2. Relative abundance of macro-invertebrate taxa (% of total numbers cumulated over exposure period, – = absent) found in leaf packs at three sites (S 1–3) in the Ver-lorenwasser during five weeks of exposure in period 1 (fragmentation) and period 2 (flow

resumption).

Period 1 Period 2

S 1 S 2 S 3 S 1 S 2 S 3

Gammarus pulex L. 4.6 4.3 0.5 – – –Asellus aquaticus L. 57.5 45.4 90 47.6 57.5 75.1Allogamus uncatus BRAUER – – – 1 0.7 0.7Glyphotaelius pellucidus RETZIUS

– – – 5.7 0.7 0.7

Platambus maculatus L. 3.4 8.6 3.4 2.9 8.9 3.5Nepa cinerea L. 1.1 0.5 – 1 0.7 0.4Erpobdella octoculata L. 2.3 3.2 0.5 2.9 1.4 2.6Helobdella stagnalis L. 1.1 18.4 0.3 1.9 0.7 3.5Glossiphonia complanata L. 4.6 2.7 0.3 1.9 0.7 3.7Grammotaulius spp. – – – 2.9 1.4 0.4Micropsectra spp. 14.9 11.4 3.7 12.4 9.6 4.6Neumora spp. – – – 1.9 0.7 0.2Psidium spp. 3.4 0.5 0.3 1.9 1 0.4Ptychoptera spp. – – – 1 0.7 0.4Culicidae 1.1 2.2 0.5 1 0.7 0.2Simulidae – – – 2.9 2.1 0.4Tipulidae – – – 0.1 0.7 0.2Phryganeidae 2.3 0.5 – 1.9 0.7 0.4Ostracoda – 2.2 – 1.9 2.1 0.4Plathelminthes 3.4 – 0.5 1.9 2.1 0.2Nemathelminthes – – – 2.9 1.4 0.7Ceratopogonidae – – – 0.1 1.8 0.4Mycetophilidae – – – 1.9 1.8 0.2Therevidae – – – 0.1 1.4 0.4SUM 100 100 100 100 100 100Total no. of taxa 12 12 10 23 23 23



051015202530

01

23

45

Shredder(no. g AFDM

-1)

S 1

(per

iod

1)S

2 (p

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S 3

(per

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1)S

1 (p

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d 2)

S2

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a)

0246

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Non-shredder(no. g AFDM

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re 3

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Non-shredder(mg g AFDM

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3.4. Leaf Mass Loss

Alder leaves appeared physically unaffected with no visual signs of shredding during each five-week exposure period. Exposed alder leaf packs lost 23 to 30% of their initial mass within the first week in both periods. This rapid initial mass loss can be attributed to leaching of readily soluble leaf compounds (Fig. 5). During the following four weeks, mass loss was slower and varied between 10 and 20% depending on period and site. Only at S 3 almost no mass loss occurred after the second week during fragmentation. Overall, leaf mass remaining did not differ between periods (ANCOVA, P = 0.17), but a significant effect of the covariates was detected. ANOVA revealed a significant effect (P < 0.05) of extreme drought conditions on leaf mass remaining during fragmentation. Decay rates calculated on a days-of-exposure basis did not differ between periods (ANCOVA, P = 0.22), but when calculated on a degree day basis, in order to compensate for temperature differences, rates were higher (ANCOVA, P < 0.05) after flow resumption than during fragmentation (Table 4).

0.0

0.5

1.0

1.5

2.0

2.5

0 1 2 3 4 5

Res

pira

tion

rate

(mg

O2

g-1le

af A

FDM

h-1

)

Site 1 (period 1) Site 2 (period 1) Site 3 (period 1)

Site 1 (period 2) Site 2 (period 2) Site 3 (period 2)a)

0

2

4

6

8

0 1 2 3 4 5


Spor

ulat

ion

rate

(con

idia

µg

leaf

AFD

M-1

h-1)

b)

Figure 4. Leaf-associated respiration (a) and fungal sporulation (b) rates during five weeks of leaf pack exposure in the Verlorenwasser. Data shown are means ± standard errors of weekly measurements at the

sampling sites (S 1–3) during period 1 or 2, respectively. AFDM = ash-free dry mass.



Table 3. Relative contribution of aquatic hyphomycete taxa (% of total numbers cumu-lated over exposure period, – = absent) to conidia production associated with exposed alder leaves at three sites (S 1–3) in the Verlorenwasser during five weeks of exposure in period 1

(fragmentation) and period 2 (flow resumption).

Period 1 Period 2

S 1 S 2 S 3 S 1 S 2 S 3

Alatospora acuminata INGOLD – – – 66.9 67 71.2Anguillospora longissima (SACC. and SYD.) INGOLD

– – – 1.1 0.5 0.2

Articulospora tetracladia INGOLD – – – 1.3 0.1 0.4Clavariopsis aquatica DE WILD. – – – 0.8 0.8 2.3Culicidospora gravida PETERSEN 31.3 50 50 13.8 13.8 12.9Diplocladiella scalaroides ARNAUD – – – 0.1 0.1 –Flagellospora curvula INGOLD 31.3 – 15.4 – – –Flagellospora fusarioides IQBAL – – – 4.0 0.5 –Heliscella stellata (INGOLD and COX) MARVANOVÁ

12.5 25 11.5 7.5 7.5 5

Lemmoniera aquatica DE WILD. 12.5 12.5 11.5 0.5 – 0.1Lemmoniera terrestris TUBAKI – – – 0.1 0.1 –Mycocentrospora acerina (HARTIG) DEIGHTON

– – – 0.5 4 1.7

Tetracladium marchalianum DE WILD.

– – – 3.0 3 4.4

Tetrachaetum elegans INGOLD – – – 0.1 1.3 1Tricladium varium JONES and STEWART

– – – 0.1 0.1 –

Unidentified 1 (tetraradiate) – – – 0.2 1.1 0.4Unidentified 2 (sigmoid < 100 μm) – – – – 0.3Unidentified 3 (sigmoid 100–120 μm)

– – – 0.2 0.3

Unidentified 4 (sigmoid > 120 μm) 12.5 12.5 11.5 0.1 – –SUM 100 100 100 100 100 100Total no. of taxa: 5 4 5 16 15 13

30

40

50

60

70

80

90

100

0 1 2 3 4 5Exposure time (weeks)

AFD

M re

mai

ning

(%)

S 1 (period 1) S 2 (period 1) S 3 (period 1)S 1 (period 2) S 2 (period 2) S 3 (period 2)

Figure 5. Mass remaining of leaf packs during five weeks of exposure in the Verlorenwasser. Data shown are means ± standard errors of weekly measurements at the sampling sites during period 1 or 2,

respectively. AFDM = ash-free dry mass.



4. Discussion


In our study, the groundwater influence within the stream varied between periods. We assume that the groundwater dominated the environmental conditions in the isolated pools during fragmentation. This assumption is corroborated by pool persistence throughout the hydrological drought period probably sustained by the inflowing groundwater into the Ver-lorenwasser. Analogously, pool persistence during drought was observed by STANLEY et al. (1997) and DAHM et al. (2003) and attributed to groundwater inflow through porous stre-ambed sediments. Additionally, flow velocity appeared extremely reduced during fragmenta-tion of the Verlorenwasser, but water flow did not cease completely leading to extremely low but measurable flow velocity. This, again, is probably caused by the inflowing groundwater into the Verlorenwasser pools and contrasts with conditions described for other intermittent streams without groundwater inflow (MOLINERO and POZO, 2004; CANHOTO and LARANJEIRA, 2007), where flow stopped completely. Furthermore, the groundwater dominance is reflected by the simultaneous occurrence of pool isolation in the Verlorenwasser and a sharp decrease of the mean daily water temperature from 20.2 to 17.4 °C, while air temperature remained high. Similar to the Verlorenwasser, LABBE and FAUSCH (2000) measured that groundwater inflow into remnant pools of a drought-affected Colorado plain stream moderated water temperature despite high air temperature. Contrary, water temperature in other fragmented streams rather followed the atmospheric temperature, and either increased or remained stable during fragmentation (BOULTON and LAKE, 1992; DAVIS et al., 2003; ACUÑA et al., 2005), presumably indicating lower or no groundwater influence of these streams during drought.

Most physicochemical parameters during fragmentation differed significantly from param-eters after flow resumption. We assume that these differences were mainly caused by dif-ferences in groundwater influence between periods. Thus, oxygen concentrations appeared extremely reduced during fragmentation probably due to the inflow of oxygen-poor ground-water. However, low oxygen levels were also measured by other investigators in fragmented streams (STANLEY et al., 1997; CARUSO, 2002) as a typical response to pool isolation and attributed to oxygen consuming heterotrophic processes stimulated by large amounts of organic matter that commonly accumulate in pools during fragmentation. Possibly, similar

Table 4. Decay rates of alder leaf packs exposed for five weeks in period 1 (fragmentation) and period 2 (flow resumption) at three sites (S 1–3) in the Verlorenwasser. Rates were either calculated on days of exposure (k) or degree-days (k') basis. Values represent means ± stand-

ard error. Coefficients of determination (R2) indicate how the data fit the regression.

Period 1 Period 2

S 1 S 2 S 3 S 1 S 2 S 3

k days–1 0.009*** ± 0.003

0.009* ± 0.003

0.002* ± 0.003

0.018*** ± 0.002

0.018*** ± 0.002

0.014*** ± 0.002

R2 0.78 0.38 0.38 0.86 0.84 0.79k' degree-days–1 0.0010***

± 0.00010.0009*

± 0.00040.0003*

± 0.00030.0021***

± 0.00030.0022***

± 0.00120.0017***

± 0.0008R2 0.80 0.39 0.33 0.80 0.84 0.77

*(P < 0.05) and ***(P < 0.001) denote significance for regression lines for breakdown rates.



processes together with oxygen-poor groundwater affected the oxygen levels in the isolated Verlorenwasser pools. Moreover, the concentrations of SRP, NH4-N, TP, and hence conduc-tivity, were higher during fragmentation than after flow resumption and could have been influences by the groundwater. Several authors measured increased nutrient concentrations in pools of fragmented streams (e.g., CARUSO, 2002; ACUÑA et al., 2004; ACUÑA et al., 2005; CANHOTO and LARANJEIRA, 2007). However, LAKE (2003) summarized that streams receive less nutrient input via runoff during drought events, but that in-stream nutrient concentra-tions may increase due to nutrient-releasing organic matter decomposition and evaporation. These processes could have also caused the higher nutrient level and conductivity in the Ver-lorenwasser during fragmentation. Only NO3–N and TN concentrations behaved contrarily and appeared strongly increased after flow resumption. Possibilities for these differences are, again, differences in quality of groundwater and surface water sources that controlled water chemistry in both periods (DEWSON et al., 2007) or differences in the symbiotic N2-fixation capacity of dominant riparian trees (e.g., Alnus glutinosa) between both periods. The latter is proved by ZAHRAN (1999), who found that if limiting factors (e.g., water stress during drought) impose limitations on a riparian host plant, the N2-fixation of a rhizobial bacterial

strain may be reduced, and hence less N2 is drained into a stream. Similar high NO3–N concentrations were obtained by DAHM et al. (2003) after a drought period of a stream and explained by aerobic processes in the unsaturated formerly anaerobic zone above the water level during drought that produced nutrients, such as nitrate, available for mobilization and transport when flow resumes. Analogously, the supra-seasonal hydrological drought in the Verlorenwasser might have decreased the general water level extremely causing mineraliza-tion of formerly bound N compounds, which could have been washed into the stream after the water level rose. However, specific measurements of N-pathways, groundwater chemis-try or the N-fixation capacity of riparian trees were not performed in the present study, but could help to explain the observed differences.

Moreover, despite differences in environmental conditions between periods, differences within the fragmentation period occurred between pools and were caused by large daily water level fluctuations (up to 10 cm). Hence, the water level in one pool during our leaf pack exposure was temporarily lowered around the streambed level creating extreme drought conditions in streambed substrates and leaf packs that were temporarily covered by only a thin water film, while in other pools the water level was permanently more than 10 cm above streambed level. Temporary water level decreases around the streambed level were also described by WILLIAMS and HYNES (1977) to occur in drought-affected streams with groundwater dominance. GRIBOVSZKI et al. (2010) gave the water consumption of riparian vegetation in temperate streams as the most important diurnal water level fluctuation-induc-ing factor, which presumably also caused the fluctuations in the Verlorenwasser.

4.2. Macro-Invertebrate Response

Leaf-associated macro-invertebrates appeared affected during fragmentation of the Ver-lorenwasser, as indicated by the extremely low abundance and biomass. Exposed alder leaf packs within the same season either of previous years without stream fragmentation of the Verlorenwasser (SCHLIEF and MUTZ, unpublished data), or of an adjacent non-fragmented stream (SCHLIEF and MUTZ, 2009) contained up to 100 times higher macro-invertebrate abundance. We assume that macro-invertebrates had been affected before the start of the study, presumably due to harmful drought-related environmental conditions. Moreover, at the beginning of our study, shredder abundance and biomass further decreased simultane-ously with pool isolation, suggesting a proceeding deterioration of environmental conditions with effects on stream macro-invertebrates. Similarly, in other drought-affected streams, a decrease in macro-invertebrate abundance during fragmentation was observed and attributed



to press disturbance by drought-related factors (BOULTON and LAKE, 1992), which were mostly water deoxygenation (MEIJERING, 1991; STANLEY et al., 1997; TOMAN and DALL, 1998) and toxic effects of leachate compounds (CHERGUI et al., 1997; CANHOTO and LARANJEIRA, 2007; SCHLIEF and MUTZ, 2007). Increased predation pressure had also been observed during stream fragmentation (MCINTOSH et al., 2002; ACUÑA et al., 2005), since the availability of spatial refuges decreases when stream volume shrinks and macro-invertebrates are trapped in pools and cannot escape predation by moving up- or downstream. It can be assumed that the shredders in the Verlorenwasser were similarly affected by these factors during fragmenta-tion since, besides be trapped in isolated pools, they were exposed to hypoxia and predatory taxa, such as Nepa cinerea, Platambus maculatus and several Hirudinea species, all of which are known to significantly reduce the densities of their prey (DAHL and GREENBERG, 1997). However, specific tests for toxic leachate compounds (e.g., phenolics) were not conducted in this study, but their presence can be expected due to the leaching of soluble compounds (NYKVIST, 1963) from freshly fallen submerged leaves during fragmentation.

Relatively high leaf-associated shredder abundance and biomass were observed in one pool on one sampling date during fragmentation of the Verlorenwasser suggesting a high variability of the community structure in time and space. Similarly, in fragmented streams of warmer climate regions, it was found that macro-invertebrate abundances temporally vary during the drying process and may increase soon after pool isolation in response to contracting habitat space, and subsequently decrease due to the deteriorating environmental conditions (PETERSON et al., 2001; ACUÑA et al., 2005). Moreover, pools of a fragmented stream can differ in the community structure of aquatic organisms (MEYERHOFF and LIND, 1987; STANLEY et al., 1997).

Macro-invertebrate abundances and biomasses after flow resumption remained at a low level. This could have been a seasonal effect with generally slow macro-invertebrate recolo-nization during the cold end of winter/start of spring period. A long lag in macro-invertebrate recolonization was also observed by BOULTON and LAKE (1992) after a drought in an Aus-tralian stream. A lag in recolonization of several flying taxa in a stream in Northern Spain was attributed to the timing of their egg hatch which did not match the season (late autumn/winter) during flow resumption (OTERMIN et al., 2002). In the Verlorenwasser, besides sea-sonal effects on recolonization, the supra-seasonal drought could have been so severe that it caused the disappearance of most drought refugia with implications for macro-invertebrate recovery.

The abundances of the dominant taxa in the Verlorenwasser were similar during fragmen-tation and flow resumption. Thus, taxa which are able to replenish their air supply above the water surface, e.g., the predators Platambus maculatus and Nepa cinerea, appeared particularly drought-tolerant. These taxa seemed able to tolerate all conditions associated with drought, such as low oxygen levels, high solute concentrations and decreasing water levels, suggesting a high resilience. This strategy for drought survival of aquatic organisms was also observed in other fragmented streams (STANLEY et al., 1997; ACUÑA et al., 2005). Moreover, the shredder A. aquaticus dominated both periods. The particularly high resilience of A. aquaticus was also documented in previous studies either by high abundances in sam-ples of drought-affected streams (IVERSEN et al., 1978; LEBERFINGER and HERRMANN, 2010) or experimentally by a high survival of A. aquaticus under different simulated drought levels (LEBERFINGER et al., 2010). The commonly low abundance of A. aquaticus in an English lowland river even increased during a drought period (EXTENCE, 1981). A high abundance of A. aquaticus in a Danish lowland stream during drought was attributed to its ability to survive in water-saturated air for a short period or to move in the streambed substratum to avoid desiccation (MOTH-IVERSEN et al., 1978). Several individuals of A. aquaticus in one pool of the Verlorenwasser were detected in the exposed leaf packs during periods of no water above the streambed level suggesting that the leaf litter served as refugia to survive periods of extreme drought conditions. Similarly, BOULTON (1989) found several stream



macro-invertebrates to use refugia, such as leaf litter or space under stones, to protect them-selves against desiccation during periods of no surface water. Other studies documented that the hyporheic zone of temporary streams can be used as an important refuge when surface flow recedes (HOSE et al., 2005). However, the Verlorenwasser profiles indicated an oxygen level in the subsurface water of around 0 mg l–1. Although it is known that A. aquaticus can survive anoxia for a certain period (HERVANT and MATHIEU, 1995), the litter coverage presumably provided more favourable conditions to survive periods of extremely reduced water level in the Verlorenwasser.

Besides A. aquaticus, other shredder species were lacking during fragmentation or disap-peared at the beginning of pool isolation in the Verlorenwasser. This could be due to the synchronism of the phenology of their life histories with the season in which the drought period occurred, or to harmful drought-related environmental conditions. Thus, trichopteran shredders were lacking in the Verlorenwasser during fragmentation, but appeared soon after flow resumption. A similar observation was obtained by OTERMIN et al. (2002) in a drought-affected stream in Northern Spain and attributed to the timing of the life histories of the Trichoptera taxa. It is known that Glyphotaelius pellucidus, one of the trichopteran shred-ders found in the Verlorenwasser, emerges in late-spring and deposits its eggs on terrestrial riparian plants closely above the water level and the egg hatch occurs when water level rises (OTTO, 1987). Probably, the emergence of the trichopteran shredders in the Verlorenwasser occurred before the drought and their eggs hatched during flow resumption. The disappear-ance of G. pulex, a shredder species that was previously abundant in macro-invertebrate samples and coexisted with A. aquaticus in the Verlorenwasser during baseflow conditions in 2004 and 2005 (SCHLIEF and MUTZ, unpublished data) and in an adjacent stream in sum-mer 2006 (SCHLIEF and MUTZ, 2009), was presumably caused by harmful drought-related environmental conditions. One possibility could be a deleterious effect induced by extremely low oxygen concentrations in isolated pools. Previous studies gave contrary results concern-ing the tolerance of G. pulex to low oxygen. Thus, laboratory experiments revealed a high mortality of G. pulex during 24 h exposure to oxygen concentrations below 2 mg l–1 (MALT-BY, 1995), whereas in other studies, G. pulex has been known to survive short-term (< 12 h) pulses of oxygen levels below 1.5 mg l–1. However, proof is needed whether G. pulex can survive hypoxic conditions during a longer period (e.g., several weeks or months). Another critical effect of hypoxia on gammarids is the stimulation of emergence from benthic refugia resulting in either downstream drift or upstream movement (HOBACK and BARNHART, 1996). During pool isolation in the Verlorenwasser, the G. pulex individuals were trapped in pools and hence possibly stranded while performing this behaviour. This could be an explana-tion for the numerous dead G. pulex individuals observed at the beginning of pool isola-tion. Moreover, G. pulex did not recolonize the Verlorenwasser even several months after flow resumption. Similarly, other investigators observed that drought events, particularly supra-seasonal droughts, may leave lingering signals in the community structure of streams (BOULTON and LAKE, 1992). Thus, as in the Verlorenwasser, G. pulex were lacking after a drought in an English chalk stream (LADLE and BASS, 1981). After an unprecedented long drought in a Danish lowland stream, the previously frequent G. pulex was eliminated and replaced by the previously lacking A. aquaticus, which established a stable population dur-ing flow resumption (MOTH-IVERSEN et al., 1978). However, whether the low oxygen level alone or in combination with other factors (e.g., high leachate concentrations) were deleteri-ous for G. pulex in the Verlorenwasser could not be answered by the present study, but a harmful effect of drought-related environmental conditions can be assumed since numerous dead G. pulex individuals were observed at the stream margins at the beginning of pool isolation.



4.3. Microbial Response

Leaf-associated microbial activity appeared strongly reduced during fragmentation of the Verlorenwasser, as reflected by extremely low leaf-associated respiration and retarded and low sporulation by aquatic fungi. Measurements with alder leaves exposed in an adjacent non-fragmented stream within the same season revealed two to three times higher respira-tion and sporulation which rapidly increased to maximum values that were four times higher than in the present study (SCHLIEF and MUTZ, 2009). Moreover, the maximum respiration during fragmentation was less than half the amount measured after flow resumption in the present study. One may assume a contrary result purely based on water temperature, since it was more than 11 °C lower after flow resumption than during fragmentation, and leaf-associated respiration usually decreases with lower temperatures (TANK et al., 1993). We assume that this discrepancy was caused by the different environmental conditions, other than temperature, of the two exposure periods that affected the microbial activity. We assume that particularly the oxygen level of the water during fragmentation, which was more than 6 mg O2 l–1 lower than after flow resumption, may have limited the oxygen supply to leaf-associated microorganisms and could explain the differences in microbial activity. Moreover, the leaf associated microbial colonization was presumably lowered dur-ing fragmentation. This assumption is supported by the extremely low and delayed sporu-lation of aquatic hyphomycetes during fragmentation, which was around 30 times lower than after flow resumption and firstly occurred after five weeks of leaf exposure. Similarly, MEDEIROS et al. (2009) found lower diversity and activity of aquatic fungi associated with alder leaves in microcosms during hypoxia compared to normoxia. Besides differences in oxygenation differences in nutrient, in particular nitrate, concentration could have supported the differences in the microbial activity between the periods. Thus, GULIS and SUBERKROPP (2003) observed higher leaf associated respiration in a nutrient-enriched than in an unaltered reach of a headwater stream. Moreover, FERREIRA et al. (2006) measured higher sporulation of aquatic hyphomycetes associated to leaf litter exposed in several nitrate enriched stream sites comparing to reference conditions. Accordingly, the microbial activity could have been stimulated by better oxygenation and higher nitrate concentration after flow resumption in the Verlorenwasser.

Aquatic hyphomycetes, obviously limited during fragmentation, appeared to rapidly rec-olonize the Verlorenwasser after flow resumption, as demonstrated by the three times higher numbers of taxa. As suggested by KRAUSS et al. (2003), the groundwater may have played a vital role as a long-term reservoir facilitating the rapid recolonization following the strong decrease in fungal communities in the surface water. However, two species, Culicidospora gravida and Heliscella stellata, had a high contribution to total spore release during both periods and seemed to tolerate conditions in the surface water during fragmentation and flow resumption.

4.4. Response of Leaf Decay Function

Although exposed leaf packs were colonized by shredders, alder leaves showed no visual signs of shredding during fragmentation, indicating that shredder-mediated decay is strongly reduced and leaf decay appeared to be predominantly microbial mediated. Similar results were obtained by LEBERFINGER et al. (2010) in a controlled laboratory experiment in con-tainers with strongly reduced water levels (1 cm above and below the sediment surface) to simulate extreme drought conditions. The experiment revealed that although shredders, such as A. aquaticus, were present, leaf decay was strongly reduced and attributed to decreased shredding activity induced directly by the reduced water level. However, the minimum water level in at least two of the pools in the Verlorenwasser was above the control level of this



experiment and oxygen levels in the containers were high due to regular water exchange. Thus, these causes for inhibition of leaf shredding presumably differ from our experiment. We assume that environmental conditions in the Verlorenwasser, such as hypoxia, are unfa-vourable for leaf shredding by macro-invertebrates during fragmentation. BJELKE (2005) found that A. aquaticus has the same leaf-shredding capacity at normoxia and hypoxia (2 mg l–1), but stopped shredding at extremely low oxygen (1 mg l–1) levels. The oxygen threshold value for A. aquaticus shredding during fragmentation in our study did not appear to be exceeded; or the combination of hypoxia with other factors (e.g., high leachate con-centrations) limited shredding activity.

A significant effect of site and extreme drought conditions on leaf mass remaining dur-ing fragmentation of the Verlorenwasser indicates that the extent of the drought conditions of a specific pool (e.g., pools with water level either permanently or temporarily above streambed level) may lead to differences in decay rates. Thus, the lowest rates in the Ver-lorenwasser with almost no leaf mass loss after week two of exposure were measured in the pool with temporarily decreased water level around streambed level. LEBERFINGER et al. (2010) found under controlled laboratory conditions that a reduction of the water level to 1 cm above the sediment surface slowed down shredder-mediated leaf breakdown compared to reference conditions with 6 cm above the sediment surface. However, LEBERFINGER et al. (2010) attributed this reduction to reduced shredder-mediated decay. Leaf decay during fragmentation in our study could not have been attributed to shredders, since no feeding signs had been observed. Thus, the extreme decrease of the water level to streambed level presumably affected the microbial-mediated decay, although the leaf associated microbial activity did not differ between the Verlorenwasser pools. Possibly, our measurements of leaf associated microbial activity that were performed in stream water filled flasks did not reflect conditions of extremely lowered water level. Studies about microbial leaf decay under extremely decreasing water levels gave variable results. While some reported a reduction (INKLEY et al., 2008), others reported an increase (BATTLE and GOLLADAY, 2001) that was attributed to promoted microbial activity through better aeration during periods of drying. However, the water level in the Verlorenwasser pool only decreased around streambed level and the leaf packs did not dry up. Studies are needed to prove the significance of those specific conditions for leaf-associated microorganisms and microbial-mediated litter decay.

When comparing both periods, daily decay rates were similar, but temperature-corrected (degree-day based) rates were lower during fragmentation than after flow resumption. Since temperature is known to be of importance in controlling leaf decay (ROWE et al., 1996) and temperature differences between both periods were high (more than 11 °C), the comparison of the temperature-corrected rates is considered appropriate and indicates a stimulation of leaf decay processes after flow resumption. This might have been caused by stimulated microbial-mediated decay as corroborated by the higher microbial activity during this period. In contrast, the contribution of shredders to leaf decay seemed to be still reduced, as indicat-ed by the lack of shredder feeding signs despite shredder presence in leaf packs. Moreover, the degree-day based decay rates after flow resumption of the Verlorenwasser were only half the amount of rates measured in an adjacent non-fragmented stream for exposed alder leaves in coarse mesh-bags with shredder access to leaves, and similar to rates in fine mesh-bags without shredder access to leaves (SCHLIEF and MUTZ, 2009). This again corroborates the assumption that decay rates after flow resumption were mainly microbial-mediated and still reduced due to limited shredder contribution to leaf decay. Presumably, the low shredder abundance alone, or in combination with a limiting effect of environmental factors (e.g., low water temperature), caused the low shredder mediated decay rates after flow resumption.

However, alder leaves are generally considered as a palatable leaf species for consumers and known to decompose faster than more refractory ones, such as oak (LECERF et al., 2007). The presence of thick oak litter layers and fewer alder leaves on the stream bottom after flow resumption may indicate that the refractory leaf species decomposed at far lower rates



than alder and hence, tend to remain for even a longer time in the stream. Measurements of decomposition rates of different leaf species during a longer exposure period (e.g., several months) in drought-affected streams could help to clarify this speculation.

5. Conclusions

The supra-seasonal hydrological drought of 2006 in the Verlorenwasser enabled the dem-onstration of possible effects of an extreme drought event for a temperate groundwater-fed lowland stream. In general, our results indicate that extreme drought events have the potential to slow down in-stream leaf decay processes during fragmentation and after flow resumption. Consequently, leaf litter can remain abundantly in a stream even several months after flow resumption. The slow decay rates during and after a supra-seasonal drought and the lack of rapid recolonization of shredders imply that extreme drought events can have long-term impacts on aquatic community structure and ecosystem processes. Globally, we suggest the providence of data about such extreme events, which are predicted to occur more frequently due to climate change, to forecast possible impacts on aquatic ecosystems.

6. Acknowledgements

We thank THOMAS WOLBURG, MICHAEL SEIDEL and MATHIAS KÜMMERLEN for their assistance in the field and lab. We also thank the Lindenberg Meteorological Observatory “Richard Aßmann Observa-tory” for the provision of regional precipitation and temperature data. The study was part of the priority program AQUASHIFT (SPP 1162, Mu 1464/2–2) supported by the German Research Foundation (DFG).

7. References

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Manuscript submitted January 12th, 2011; revised August 1st, 2011; accepted September 5th, 2011

leaf decay processes during and after a supra-seasonal hydrological drought in a temperate lowland...

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