Introduction

Dissolved organic matter (DOM) is the major pool of reduced carbon (C) in transport from the continents to the ocean (Volk et al. 1997) and represents the primary substrate for microbial communities in freshwaters (Benner 2003). The amount and composition of DOM in rivers is the combined result of its different sources and in-stream transformation processes such as biodegradation, photodegradation, or flocculation (Aitkenhead-Peterson et al. 2003; Sinsabaugh and Foreman 2003; Sachse et al. 2005). Depending on its source, DOM can be classified either as autochthonous, when produced within the aquatic system itself, or allochthonous when it derives from terrestrial ecosystems. Autochthonous DOM is generally assumed to be rich in biologically reactive components due to its inherent properties (e.g. low molecular weight, high nitrogen content; Bertilsson and Jones 2003), whereas terrestrial DOM has been traditionally considered as an heterogeneous mixture of contemporary and aged compounds (Raymond and Bauer 2001; Butman et al. 2012) that is recalcitrant and contributes little to freshwater ecosystem metabolism (Wetzel 1983). However, the high prevalence of net heterotrophy and carbon dioxide (CO2) supersaturation in freshwaters implies that terrestrially derived C is significantly contributing to ecosystem respiration (Cole and Caraco 2001; Duarte and Prairie 2005; Battin et al. 2008) and subsidizing river food webs (Caraco et al. 2010), even if this C is ancient (McCallister and Del Giorgio 2012; Fellman et al. 2015). Therefore, a comprehensive accounting not only of the sources but also the biodegradability of DOM is paramount to elucidate the role of river networks on the continental C cycle (Cole et al. 2007; Battin et al. 2008).

Because most riverine DOM originates from terrestrial ecosystems (Wetzel 1983; Cole and Caraco 2001), its specific sources are highly dependent on the hydrological connection between the river channel and both the riparian and groundwater systems (Aitkenhead-Peterson et al. 2003). Nevertheless, how the lack of hydrological connectivity affects the sources and biodegradation of DOM in rivers has been rarely explored. This situation often takes place in arid and semiarid regions, where river networks are subject to a dynamic hydrological regime that usually involves drought episodes (Gasith and Resh 1999). During these periods, lowering water tables reduce the hydrological connectivity between the river and its catchment as well as along the river flow path (i.e. longitudinal fragmentation; Bernal et al. 2013). In addition, such loss of connectivity is enhanced by a myriad of man-made flow discontinuities such as weirs and impoundments, which are widespread in these regions (Döll et al. 2009). In this sense, drought turns river networks into a fragmented and heterogeneous landscape dominated by slow running waters, isolated water pools and weirs, where the hydrological connectivity with terrestrial ecosystems is reduced to deeper groundwater inputs (Dahm et al. 2003).

During drought, the prevalence of more autochthonous DOM has been reported in semiarid temporary streams in comparison to other hydrological situations (Vazquez et al. 2011; Ylla et al. 2011; von Schiller et al. 2015). Particularly, the last stages of drought when the river network can be highly fragmented have been described as hot moments of autochthonous DOM production (von Schiller et al. 2015; Vazquez et al. 2015). Recent findings also highlight great spatial heterogeneity in the quantity and composition of the riverine DOM during these periods, with isolated pools standing out as distinct individual ecosystems due to the longitudinal fragmentation (Fellman et al. 2011; Vazquez et al. 2011; Siebers et al. 2015). However, the specific DOM sources and in-stream processes that provide this heterogeneity are still unknown. Moreover, such spatial heterogeneity in DOM composition may in turn imply variable DOM biodegradability along the river (e.g., Vazquez et al. 2011). Hence, DOM of different nature (e.g., source and radiocarbon age) may be predominantly respired in each different aquatic environment.

Here we aimed at identifying the different sources of DOM and the DOM pools that were respired along a fragmented river under drought conditions. To do so, we characterized the DOM present in the different aquatic environments of a Mediterranean river during summer drought (isolated pools, running waters, and impounded waters in weirs) using a multi-faceted approach that combines absorbance-fluorescence spectroscopy, size-exclusion chromatography (SEC), biodegradation assays, and stable and radiocarbon isotope signatures of both organic and inorganic dissolved C (DOC and DIC, respectively). We expected (i) to identify heterogeneous DOM sources along the river, (ii) a partial disconnection from terrestrial DOM sources, and (iii) respiration of different DOM pools at each aquatic environment.

Methods

Study site

The Fluvià is a Mediterranean-type river located in the NE of the Iberian Peninsula (Fig. 1). Its main stem is 97 km long, with a mean flow of 3.6 m3 s−1 in the basin outlet (data from 2004 to 2014; Catalan Water Agency). Its river network drains a 990 km2 catchment covered by mixed forests (79 %) and agricultural (18 %) and urban (3 %) areas. The catchment geology is calcareous, with some areas (<15 %) of siliceous materials (Cartographic and Geological Institute of Catalonia, 2006). Climate is typically Mediterranean, with air temperatures ranging from 6 to 26 °C and a mean annual precipitation of 660 mm (data from 2004 to 2014; Catalan Water Agency) mainly occurring in autumn and spring with occasional summer storms. The hydrologic regime follows the precipitation pattern. During summer drought, the combination of scarce precipitation and high evapotranspiration promotes a drying pattern of the river network that begins in headwaters and advances downstream.

Fig. 1
figure 1

Location of the sampling sites along the study reach. Dashed lines symbolize reaches that were dry during the sampling period. The upper inset shows the location of the Fluvià catchment in SW Europe. The lower inset shows the drainage area associated to the study reach within the Fluvià catchment (NE of the Iberian Peninsula)

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We focused this study on a 7-km upland reach of the Fluvià river main stem affected by drought during summer (Fig. 1). The study reach is surrounded by a dense deciduous riparian forest and croplands, with no influence of industrial or wastewater treatment activities. During the study period, the reach included upstream isolated pools, 5.5 km of running waters, and the first weir on the main stem. We collected surface water from seven sites along the study reach representing these different aquatic environments. The selected sampling sites were two upstream isolated pools (“Pool1” and “Pool2”), the seepage point where groundwater appeared and surface water started to flow (“Seepage”), two downstream running water sites (“Flow1” and “Flow2”), and two sampling points in the weir: one just after the river water mixes with the lentic water body and the other near the dam (“Tail” and “Dam”, respectively; Fig. 1). Isolated pools were 500 m away from each other and were located in shaded areas under riparian canopy. Both pools contained several patches of accumulated tree leaves. However, whereas tree leaves accounted for 75 % of the Pool1 area, in Pool2 tree leaves only covered 10 %. Discharge progressively increased from Seepage to downstream running water sites, with no visible sudden changes along the study reach (Casas-Ruiz et al., personal observation). Discharge in Flow2 was 0.122 m3 s−1 (Sontek, YSI, USA), with a mean water velocity of 0.07 m s−1, and a width and mean water depth of 5.1 and 0.33 m, respectively. The substratum in both Flow1 and Flow2 was covered by epilithic and epipsammic biofilms. The area, volume and water residence time of the weir were 1.24 hm2, 0.0174 hm3 and 1.65 days, respectively. The water depth in the weir was 0.6 m in Tail and 3.9 m in Dam.

Field sampling

Sampling was carried out in late summer (September the 1st, 2013) during a seasonal drought episode. Water samples for DOC and DIC concentrations (three samples per site) were collected with 60 mL syringes, immediately filtered in the field through pre-combusted and pre-rinsed 0.7 µm glass fiber filters (GF/F; Whatman, UK), and placed without headspace into polyethylene plastic bottles after rinsing several times with filtered stream water. Samples for inorganic and organic C isotopes were collected without previous filtration in pre-cleaned (0.1 N HCl and Milli-Q) and combusted (4 h, 500 °C) 2 L glass bottles, capped with solid ground glass joints. Inorganic C samples were stabilized in the field with mercury chloride solution. Samples for spectroscopic measurements and SEC were filtered in the field through pre-rinsed 0.2 µm nylon filters (Whatman, UK) and placed into 11 mL polypropylene plastic containers (three samples per site). Water samples were kept in cool and dark conditions during transport to laboratory. Samples for DOC and DIC concentrations and spectroscopic measurements remained refrigerated (<4 °C) in dark conditions, whereas those for SEC analyses were kept frozen until analysis. In addition, two samples of fresh leaves were collected in different locations along the study reach to estimate the 14C signature of a recently produced terrestrial DOM source. We also used portable probes to measure in situ water temperature, conductivity, pH (WTW, Germany), and dissolved oxygen (DO; YSI ProODO Handheld, Ohio, USA) in each sampling site.

DOM characterization

We used DOC concentration as a quantitative indicator of DOM. All samples were analyzed for DOC concentration within one day following sample collection by high-temperature catalytic oxidation on a Shimadzu TOC-V CSH analyzer (Shimadzu Corporation, Japan).

Absorbance-fluorescence spectroscopy

UV–Vis absorbance spectra (190–800 nm) were measured on an Agilent 8453 diode array spectrophotometer (Agilent Technologies, Germany) using a 1-cm quartz cuvette. Milli-Q water was used as a blank. The average sample absorbance between 700 and 800 nm was subtracted from the spectrum in order to correct for offsets due to instrument baseline drift (Green and Blough 1994). We calculated the specific UV absorbance at 254 nm (SUVA254, L mg−1 m−1), an indicator of aromatic C content (Weishaar et al. 2003), by dividing sample absorbance at 254 nm by DOC concentration and cell length.

Excitation-Emission matrices (EEMs) were obtained using a fluorescence spectrophotometer (F-7000, Hitachi, Japan) with a 1-cm quartz cuvette. EEMs were acquired by measuring fluorescence intensity across the excitation range set from 200 to 449 nm (3 nm increments) and the emission range set from 250 to 598 nm (3 nm increments). Excitation and emission slit widths were set to 5 nm. Water samples were allowed to warm to room temperature prior to measurements. All EEMs were read in ratio mode, whereby sample and lamp reference signal were corrected for instrument-specific biases after calibrating the instrument with Rhodamine B solution and the light diffuser and filter set provided by the manufacturer. EEMs were blank subtracted using the EEM of Milli-Q water collected every ten samples. Spectra were corrected for inner filter effects using UV–Visible absorbance spectra (Lakowicz 2006) without requiring previous dilution according to Kothawala et al. (2013). The integral of the Raman scatter peak of water blanks was used for intensity calibration following the procedure described in Lawaetz and Stedmon (2009). EEMs were corrected and analyzed using MATLAB 2012a (Mathworks, MA, USA) software. We decided to interpret fluorescence data descriptively rather than by using any dimension reduction statistical approach (e.g., Zhang et al. 2014) because of the small size of our data set.

The intensity of the main fluorescence peaks related with DOM characterization (A, M, C, B, T; Coble 1996; Parlanti et al. 2000) was obtained from the EEMs. For each peak we kept the maximum fluorescence within the EEM regions defined in Fellman et al. (2010). Briefly, peaks A and C exhibit emission at long wavelengths and are related to aromatic humic-like fluorophores that derive from terrestrial sources. Peak M is also related to humic-like substances, although it exhibits emission at shorter wavelengths and is considered to be less aromatic and likely of lower molecular weight. Conversely, peaks B and T are related to tyrosine-like and tryptophan-like substances respectively, and have been associated to protein-like materials of different origin that are usually related to DOM lability (e.g., Balcarczyk et al. 2009; Guillemette and del Giorgio 2011). Fluorescence peak intensities are reported as percent contribution of each peak to the sum of fluorescence intensity by all peaks.

We also calculated the following spectral indices: the humification index (HIX), and the biological index (BIX). HIX was calculated as the ratio between the peak area under the emission spectra 435–480 nm and 300–345 nm, at an excitation wavelength of 254 nm (Zsolnay et al. 1999). HIX values increase with the extent of humification. BIX was calculated by dividing the fluorescence intensity emitted at 380 nm by that at 430 nm for an excitation of 310 nm (Huguet et al. 2009). Higher BIX values correspond to a higher autochthonous origin and to the presence of freshly released DOM (Huguet et al. 2009). We discarded the use of the fluorescence index (FI) because it showed a rather erratic behaviour in our samples, as previously reported in other studies (Korak et al. 2014).

Size-exclusion chromatography

The DOM size fractions were characterized by liquid chromatography. Briefly, the chromatographable DOC portion of the filtered samples passed through a size-exclusion column packed with resin (Toyopearl HW 50S, volume of 250 × 20 mm). Phosphate buffer (0,029 mol/L, pH 6,5) was used as eluent in a flow rate of 1.1 ml min−1. DOC was detected by infrared absorbance of CO2 after UV oxidation at 185 nm in a cylindrical UV thin-film reactor (Graentzel-reactor). Fractions were identified by using fulvic and humic acid standards from the International Humic Substances Society (IHSS). In brief, the chromatograms were deconvoluted and apparent size fractions assigned based on the retention time and column affinity obtained for the IHSS standards. Fraction assignment was done by using the software FIFFIKUS (DOC-Labor Huber, Karlsruhe). We separated DOM into three different fractions: (i) high molecular weight substances (HMWS), including polysaccharides and to a lesser extent proteins or amino sugars, (ii) humic or humic-like substances (HS), including building blocks (lower molecular weight breakdown products of HS), and (iii) low molecular weight substances (LMWS), which includes low molecular weight acids and neutrals such as sugars and aminoacids. According to Her et al. (2002) HMWS corresponded to a molecular weight of 7 kDa or higher, HS to a range of 7–0.4 kDa, and LMWS to less than 0.4 kDa. We report each fraction as the percent contribution to total DOC.

Biodegradation assays

Biodegradable DOC (BDOC) was measured in dark incubations (Amon and Benner 1996; Guillemette and del Giorgio 2011). Briefly, we incubated 60 mL of filtered water (pre-rinsed 0.2 µm nylon membranes, Whatman, UK) with an inoculum added in a 1:10 proportion consisting in water from the same site filtered by Whatman GF/F. Samples were replicated three times and incubated in the dark in an incubation chamber at 20 °C for 65 days. BDOC was calculated as the difference in DOC prior and after the incubation.

Isotopic signatures

The δ2H and δ18O ratios of water samples were measured in the Colorado Plateau Stable Isotope Laboratory, Flagstaff, USA. For analysis of 13C in DOC (13CDOC) and DIC (13CDIC), separate water samples of 10 mL volume were stabilized with 10 µL HgCl2 and sent to the Colorado Plateau Laboratory for conventional isotopic ratio mass spectrometry (MS). Samples for 14CDOC and 14CDIC were processed and measured as described in Tittel et al. (2013). Briefly, DOC was filtered in the lab within 24 h after collection (pre-combusted Whatman GF/F) and then enriched at 60 °C by evaporation with UHP nitrogen. Inorganic C was outstripped after acidification and precipitated in barium hydroxide solution. 14CDOC and 14CDIC processed samples were measured by accelerator mass spectrometry (AMS) at the Leibniz-Laboratory for Radiometric Dating and Isotope Research in Kiel, Germany. Unfortunately, only three ∆14CDOC signatures (Pool1, Pool2, and Flow1) could be successfully determined. Flow2, Tail and Dam samples were lost at graphitization in the lab, while Seepage did not contain enough C. Data for stable isotopes is reported in standard δ notation (‰) relative to Vienna Standard Mean Ocean Water for δ2H and δ18O, and standard Vienna Pee Dee Belemnite for 13C. 14C data is reported in ∆ notation with respect to oxalic acid II standard and corrected for isotope fractionation (Stuiver and Polach 1977). Inorganic C isotopic signatures were measured on unfiltered samples; however, we refer to them as 13CDIC and 14CDIC throughout the manuscript because data from previous samplings indicate that particulate inorganic C represents less than 1 % of total inorganic C in the study reach (Casas-Ruiz, unpublished data).

Influence of CO2 outgassing on δ 13CDIC

We evaluated the potential drift in δ 13CDIC values due to gas exchange along the running water reaches. To do this, we estimated the DIC evaded as CO2 between consecutive stations. Then, we solved a simple isotope mixing model to estimate the downstream δ13CDIC using the observed upstream δ13CDIC and the mass of DIC evaded as CO2.

Based on the Fick’s First Law of diffusion, the gas exchange at the water–air interface is driven by the reaeration coefficient and the gas concentration gradient. Hence, we calculated the DIC change within the reaches as

$$\frac{\partial DIC}{\partial t} = k\,z^{ - 1} \Delta {\text{CO}}_{2}$$
(1)

where \(\frac{\partial DIC}{\partial t}\) is the variation in DIC concentration with time (mol L−1 s−1), k is the gas transfer velocity (m s−1), z is the water depth (m) and ∆CO2 is the gradient between the concentration of CO2 in the water and in the atmosphere (mol L−1). The physical parameter k was obtained by applying the night-time regression method (Hornberger and Kelly 1975) on dissolved oxygen data measured every 10 min in Flow2 (mean of 4 nights). Water depth z was calculated as the mean depth from Seepage to Tail using an HEC-RAS model fed with the discharge in Flow2 and reach cross sections measured every ca. 100 m. Initial ∆CO2 was calculated using the measured DIC and pH of each reach upstream station, assuming an atmospheric CO2 partial pressure of 390 ppmv. Then, we iteratively computed ∆CO2 and DIC concentrations implementing Eq. 1 in the CO2SYS software (Lewis and Wallace 1998). The final DIC concentration in the downstream station was the DIC concentration after a time lapse equal to the travel time of the water in the reach. Once the DIC evasion between stations was obtained, we applied a simple isotope mixing model to calculate the resulting downstream δ13CDIC. Since bicarbonate was the predominant DIC fraction (>90 %) and there were no major changes in pH along the running water reaches (Table 1), we assumed an enrichment factor between DIC and CO2(g) of −9 ‰ at 15 °C (Zhang et al. 1995). The comparison between the resulting δ13CDIC due to CO2 outgassing alone and the measured δ13CDIC value in the downstream station informs about the relevance of outgassing versus other potential processes on the increasing δ13CDIC values.

Table 1 Water physicochemical parameters in the sampling sites
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Data treatment and analysis

We report data in this study as the mean and standard deviation of each variable based upon three real replicates. In the case of isotope signatures and SEC fractions, values are reported as the mean and deviation (0.5 range) of two real replicates for SEC and δ 13C, and of two analytical replicates for ∆14C. We applied Spearman correlation tests to analyze the relationship between variables. All statistical analyses were run in R 3.1.3 (R Core Team 2015).

Results

Water physicochemistry

River water δ2H and δ18O values (Table 1) ranged from −41.8 to −27.5 ‰ and from −6.8 to −5.1 ‰, respectively, and divided the study sites in two different groups: (i) the isolated pools and Seepage, and (ii) the running water sites and the weir (Fig. 2). All water samples showed very low deviation from the Local Meteoric Water Line of the Montseny area (40 km from the study area; Neal et al. 1992). No significant evidences of evaporation during soil infiltration and transport to the river channel or within the river were observed.

Fig. 2
figure 2

Relationship between δ18O and δ2H for the water samples. The global meteoric water line (GMWL) and the local meteoric water line of the Montseny area (LMWL; Neal et al. 1992) are shown. Error bars symbolize the 0.5 range of duplicate samples. Note that in some cases error bars are behind the symbol. V-SMOW; Vienna Standard Mean Ocean Water

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Water temperature was relatively stable along the study reach, with the lowest value in Flow1 (15.1 °C) and the highest in Tail (16.9 °C; Table 1). The pH varied from 7.2 to 8.2, showing lowest values in isolated pools and an increasing trend from Seepage to Dam. DO saturation widely varied across sites (from 3 % in Pool2 to 112 % in Flow2), and showed an increasing trend from Seepage to Flow2. Conductivity ranged from 336 to 724 µS cm−1, with the lowest values in isolated pools and the highest value in Flow1. DIC concentrations ranged from 32.3 to 75.5 mg C L−1, and followed a similar pattern to that of conductivity.

DOM quantity and composition

DOM quantity was highly variable along the study reach. DOC concentration ranged from 1.0 to 4.2 mg C L−1, with the highest values found in the two isolated pools and the lowest in Flow2 (Table 2). Similarly, the qualitative characterization of DOM by absorbance-fluorescence spectroscopy indicated large variation in DOM composition across sites (Table 2). SUVA254 varied from 2.0 in Flow2 to 3.0 L mg C−1 m−1 in Pool2 and Seepage, corresponding to an aromatic content of ~16 to 23 % according to Weishaar et al. (2003). HIX values widely ranged from 4.0 to 8.3, presenting the lowest values in running waters (Flow1 and Flow2) and the highest value in the tail of the weir. BIX ranged from 0.65 in Tail to 0.79 in Flow1, showing an opposite pattern to that of HIX (Fig. 3). All water samples were dominated by humic-like fluorescence, with peak A representing more than 40 % of the fluorescence derived from the combination of the five peaks in all sites. Contribution of peaks A, M, and C was relatively homogeneous across sites, showing a coefficient of variation of 5, 2, and 6 % respectively. Conversely, protein-like peaks T and B showed relatively high heterogeneity (coefficient of variation of 19 and 25 %, respectively), with highest values in Flow1 and Flow2 and lowest in Tail.

Table 2 Dissolved organic carbon concentration and absorbance-fluorescence descriptors of DOM composition across sites (Mean ± SD)
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Fig. 3
figure 3

Relationship between biological index (BIX) and humification index (HIX) across sampling sites. The continuous line follows the longitudinal flow direction from the Seepage site to the weir. Note that the position of Pool2 over the line is accidental

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SEC analyses indicate that LMWS was the dominant fraction in the running water sites and in Dam (Table 3). Conversely, isolated pools, Seepage, and Tail were dominated by HS. Only isolated pools showed relatively high contributions of HMWS (7.3 and 7.6 % in Pool1 and Pool2, respectively). The percentage of HS correlated positively with SUVA254 (r s  = 0.79, p < 0.05) and HIX (r s  = 0.71, p = 0.09). A more detailed comparison between SEC and absorbance-fluorescence results is shown in Table 4.

Table 3 Size fractions of dissolved organic matter and proportion of biodegradable dissolved organic carbon (BDOC) in the sampling sites
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Table 4 Spearman correlations between absorbance-fluorescence descriptors and size-exclusion chromatography fractions
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The percentage of BDOC ranged from 11 to 42 % (Table 3). DOM in running waters showed the highest biodegradability, while Tail had the least bioavailable DOM along the study reach. Overall, the percentage of BDOC decreased with increasing HIX and the percentage of HS (r s  = −0.78, p < 0.05 and r s  = −0.82, p < 0.05, respectively; Fig. 4a) as well as SUVA254 (r s  = −0.85, p < 0.05). Conversely, there was a positive correlation between the percentage of BDOC and the percentage of LMWS (r s  = 0.89, p < 0.05; Fig. 4b) and a marginally significant relationship between the percentage of BDOC and protein-like peak T (r s  = 0.75, p = 0.07; Fig. 4c).

Fig. 4
figure 4

Relationship between the percentage of biodegradable dissolved organic carbon (BDOC) and a the humification index (HIX), b the percentage of low molecular weight substances (LMWS), and c the percentage of protein-like Peak T. Results from Spearman correlation tests are shown

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Isotopic signatures

C isotopic signatures of DOC and DIC are given in Table 5. ∆14CDOC ranged from −108 to −158 ‰, corresponding to a conventional radiocarbon age (CRA) of 855 to 1318 year B.P. ∆14CDOC of riparian leaves (21 ‰) corresponded to a modern origin (i.e., produced after 1950). δ13CDOC values showed little variability along the study reach, with values ranging from −28.2  ‰ in Seepage to −26.0 ‰ in Pool1. ∆14CDIC ranged from −1 ‰ in Pool1 to −52 ‰ in Flow2 and Tail, corresponding to CRA of Modern to 370 year B.P. δ13CDIC values varied from -13.3 to -9.6 ‰ (Fig. 5), showing a 13C enrichment as water flowed downstream, and a positive relationship with DO saturation (r s  = 0.77, p < 0.05). Assuming CO2 outgassing as the only process governing δ13CDIC along the running water sites, we estimated hypothetical δ13CDIC values of −11.5 and −11.1 ‰ in Flow1 and Flow2, respectively (see "Methods" for calculation details). The comparison of these values with the measured ones (Table 5) suggests that CO2 outgassing might explain about 60 % of the overall δ13CDIC increase along the running water sites.

Table 5 Stable and radiocarbon isotopic signatures of dissolved organic carbon (DOC) and dissolved inorganic carbon (DIC) across the study sites
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Fig. 5
figure 5

Stable and radiocarbon isotopic signatures of dissolved inorganic carbon (DIC) across sites (a) and projected isotopic changes for processes affecting DIC in our study case (b). The grey square in b represents the range of δ13CDIC and ∆14CDIC in the flow-connected sites (i.e., from Seepage to Dam). Potential sources of DIC include degradation of tree leaves (δ13C ca. −33 ‰, ∆14C 21 ‰) and degradation of old river DOC (assumed similar to DOC in Flow1, i.e., δ13C −28 ‰ and ∆14C −158 ‰). Carbonate dissolution was not considered a potential DIC source since high and increasing pH values were found along the study reach. Error bars in a represent the 0.5 range of duplicate samples for δ13C and 0.5 range of duplicate analyses for ∆14C

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Discussion

The lack of hydrological connectivity determined the different sources of DOM as well as its degradation along the river. The old DOC measured in the study reach (855 to 1318 year B.P.) suggests that groundwater from relatively deep soil horizons was feeding the channel (Trumbore and Schiff 1992; Neff et al. 2006; Butman et al. 2007), likely caused by hydrological disconnection of the river from the riparian zone and the shallow subsurface flow (Dahm et al. 2003; Lake 2003). This lateral flow disconnection influenced the DOM sources along the study reach, and the longitudinal fragmentation promoted by flow cessation provided distinct individual environments in terms of physicochemistry and DOM properties. As a result, we found highly contrasting values of DOM descriptors and physicochemical parameters along just seven kilometres of river reach. For instance, the percentage of BDOC spanned from 11 to 42 %, a range that is comparable to or higher than the range in BDOC found across different rivers (del Giorgio and Davis 2003; Balcarczyk et al. 2009; Fellman et al. 2009, 2014). Such heterogeneity is consistent with previous studies in temporary rivers from Mediterranean (Vazquez et al. 2011; von Schiller et al. 2015) and other semiarid regions (Fellman et al. 2011; Siebers et al. 2015), where an enhanced variability of DOM composition as well as water physicochemistry occurs during hydrological fragmentation.

The physicochemical parameters measured along the river suggest that groundwater with different characteristics than Seepage contributed to the river flow (e.g., note the different δ2H, δ18O and Cond between Seepage and Flow1, Table 1). In this sense, part of the observed heterogeneity in DOM and physicochemical parameters could derive from the mixing with non-sampled groundwater. Nevertheless, the small variation in δ2H and δ18O from Flow1 to Dam (Table 1, Fig. 2) indicates that subsurface inputs promoting the perceived flow increase along the river derived all from the same source. Hence, in-stream processes might have played a relevant role on the observed relative changes in DOM and water chemistry between these sites.

Spatial discontinuities in DOM sources along the river

One of the more conspicuous heterogeneities in our data set emerges from the rapid increase in protein-like, low molecular weight DOM along the running water sites, which suggests an increasing influence of algal/bacterial DOM sources (Bertilsson and Jones 2003). Accordingly, the position of Flow1 and Flow2 in the BIX-HIX plane (Fig. 3) shows an enrichment of fresh and low-humified DOM from the Seepage site to downstream running waters, also indicated by a decrease in SUVA254. These results are in line with recent findings in Mediterranean temporary rivers, where a decrease in the proportion of terrestrial versus autochthonous DOM has been described during the contraction and fragmentation phases of drought (Vazquez et al. 2011; Ylla et al. 2011; von Schiller et al. 2015). However, we cannot identify the ultimate origin of this algal/bacterial DOM, because both autochthonous sources as well as groundwater DOM inputs between Seepage and Flow1 with a strong microbial signal (e.g., Birdwell and Engel 2010) could explain the observed shift in DOM composition.

In any case other evidences suggest that autochthonous DOM contributions, algal primary production in particular, were significant along these sites. The increase in δ13CDIC from Flow1 to Flow2 suggests that in-stream processes affecting the DIC isotopic signal took place within the running water sites (Fig. 5). Our rough estimation of the CO2 evaded along the running water sites indicates that CO2 outgassing could explain up to 60 % of the observed δ13CDIC increase, so this process alone cannot explain the overall δ13CDIC trend. Consequently, considering that carbonates dissolution is unlikely given the high and increasing pH (Table 1), photosynthesis must have been a relevant process in the running water sites, also supported by oxygen increasing above saturation in Flow2 (112 %).

The most contrasting sites in terms of DOM properties were, however, the upstream isolated pools, where we found the highest DOC concentrations. The DOM fluorescence properties and molecular weight fractions indicate that the pools were dominated by terrestrially derived, humic compounds with lower contributions of LMWS in comparison to the flow-connected sites. Such a terrestrial character is likely related to the leaching and degradation of the accumulated riparian tree leaves (see sites description), which is consistent with their relatively high proportion of HMWS (Meyer et al. 1998; Francis and Sheldon 2002). Also, the relatively young DOC radiocarbon ages compared to that in Flow1 (Table 5) are in agreement with contributions of DOC by degradation of tree leaves (δ13CPOC −33 ‰, ∆14CPOC 21 ‰). Indeed, tree leaves and detritus usually accumulate in isolated pools due to lack of advection (Acuña et al. 2005; Bond et al. 2008; von Schiller et al. 2011), causing high DOC concentrations and fuelling respiration (Sabater and Tockner 2010). In our case, the lower δ13CDIC values measured in the pools in comparison to the flow-connected sites accordingly suggest high respiration activity (Fig. 5), also indicated by the very low DO levels (2.7 and 17.4 % in Pool1 and 2, respectively). Therefore, our results suggest that leaf degradation may be the main DOM source in the pools. Nevertheless, the hydrological isolation of the pools makes them independent systems that are highly influenced by their particular local conditions. Therefore, our observations would probably differ under different situations such as less canopy cover and so less leaf inputs and higher light availability, whereby algal DOM sources may increase (Dahm et al. 2003; Fellman et al. 2011).

The tail of the weir constitutes another clear discontinuity along the study reach. The BIX-HIX plane (Fig. 3) displays Tail in the opposite corner of the immediately upstream-located running waters, which is consistent with the highest SUVA254 and percentage of HS along the study reach. This, in combination with low BDOC (11 %), indicates a highly degraded DOM in this site. The tail of weirs and impoundments where river water mixes with the lentic water body can be a high degradation environment (Šimek et al. 2010), strongly influenced by the activity of the accumulated sediment (Cardoso et al. 2013). In fact, fluxes from the sediment to the water column can be highly influential on the river chemistry during low flow periods (Mosley 2015). Hence, based on DOM composition data, we suggest both intense degradation of fresh DOM and release of highly degraded DOM from the sediment to be responsible for the humified character in this particular site. However, this interpretation is not clearly supported by physicochemical data (e.g., there is no marked δ13CDIC depletion in Tail; Fig. 5).

Spatial discontinuities in DOM degradation along the river

The relationships between BDOC and the optical and chemical properties of DOM indicate a microbial preference for protein-like, fresh, low molecular weight DOM (Fig. 4). These results pose autochthonous DOM as the most bioavailable DOM, and suggest that it may cycle rapidly within the river. However, because of the experimental ex situ approach we used, BDOC values provide information about the potential bioavailability of DOM rather than its actual in situ degradation. Therefore, in order to further recognize the DOM pools that were being degraded in each aquatic environment, we compared the variation in DIC isotopic signatures among sites (Fig. 5).

Mineralization-derived CO2 reflects the isotopic signal of the organic matter source. Hence, since we identified leaf leachates, algal release, and groundwater inputs as the principal DOM sources along the study reach, we considered them as the end-members whose degradation could modify the riverine ∆14CDIC. Following this approach and departing from a ∆14CDIC range of −34 to −52  ‰ in the flow-connected sites, inputs of CO2 resulting from the in situ degradation of tree leaves (∆14CPOC = 21 ‰) along the river would increase water ∆14CDIC values (Fig. 5b). Differently, inputs of CO2 due to the degradation of algal DOM should not greatly affect ∆14CDIC values, because ∆14C of photosynthetically derived DOC reflects that of the DIC available for C fixation. Finally, assuming the old DOC measured in Flow1 (∆14CDOC = −158 ‰; Table 5) derives from groundwater inputs, degradation of groundwater DOM within the river should decrease ∆14CDIC values.

Two distinct patterns of ∆14CDIC variation were therefore observed along the study reach. On the one hand, leaf degradation appeared to be a relevant process in isolated pools, as indicated by the higher ∆14CDIC in comparison to the flow-connected sites (Fig. 5). This is in agreement with their identification as high respiration environments (see above), and at the same time supports that the HMWS and the high DOC concentrations measured in the pools derived from the degradation of leaves. Consistently, riparian litterfall leachates have been previously suggested as a labile DOM source to rivers, especially its highest molecular weight fractions (McArthur and Richardson 2002; Romaní et al. 2006; Vazquez et al. 2015). Noteworthy, the increase in ∆14CDIC from Tail to Dam suggests that leaf leachate degradation also occurred in the weir (Fig. 5), overall emphasizing the role of tree leaves from overhanging canopies as a relevant DOM source as well as an energy substrate for aquatic heterotrophs in sites with high water residence time.

On the other hand, degradation of groundwater old DOM likely occurred within the river, as indicated by the ∆14CDIC depletion from Seepage to Tail (Fig. 5). This result is consistent with recent evidence for the respiration of aged DOM in glacier (Singer et al. 2012), boreal (McCallister and Del Giorgio 2012) and tropical ecosystems (Fellman et al. 2014), suggesting a biogeochemical link between past C fixation and contemporary river metabolism. Nevertheless, we cannot discern whether the ∆14CDIC decrease in our reach results from (i) degradation of significant amounts of moderately aged DOM, or (ii) from sporadic degradation of strongly 14C-depleted DOM. Also, we cannot fully discard that inputs of groundwater old DIC contributed to the observed ∆14CDIC depletion along the river. In any case, the positive correlations found between the percentage of BDOC and protein-like, fresh, low molecular weight DOM (Fig. 4) suggest a microbial preference for the algal/bacterial DOM measured in the running water sites. Thus, both groundwater millennial DOM and autochthonous DOM could be simultaneously degraded along the running water sites.

Conclusions

The lack of hydrological connectivity caused by drought implied a spatial discontinuity in the DOM sources along the study reach, determining at the same time the DOM pool that was being respired in each aquatic environment. The partial isolation of the river from its catchment promoted a relevant influence of autochthonous DOM sources. However, although terrestrial inputs from the riparian zone can be notably reduced, tree leaves from overhanging canopies stood out as an important bioavailable DOM source, especially in sites where water residence time was high (isolated pools and weirs). Groundwater millennial DOM (>1300 year B.P.) seemed to be respired within the river, suggesting a biogeochemical link between past C fixation and contemporary river metabolism. Although our results are inherently subject to regional and local factors that may widely differ between regions, rivers, and even reaches, they highlight a highly heterogeneous DOM pool during drought, with special emphasis on the role of local conditions (e.g., differences in light and riparian leaves availability) that may hinder any spatial upscaling.