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All issues No 424 (2023) Knowl. Manag. Aquat. Ecosyst., 424 (2023) 13 Full HTML
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Knowl. Manag. Aquat. Ecosyst.
Number 424, 2023
Biological conservation, ecosystems restoration and ecological engineering
Article Number 13
Number of page(s) 9
DOI https://doi.org/10.1051/kmae/2023010
Published online 17 May 2023

© U. Karakuş et al., Published by EDP Sciences 2023

Licence Creative CommonsThis is an Open Access article distributed under the terms of the Creative Commons Attribution License CC-BY-ND (https://creativecommons.org/licenses/by-nd/4.0/), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited. If you remix, transform, or build upon the material, you may not distribute the modified material.

1 Introduction

Invasive species can cause wide ranging detrimental impacts in their receiving ecosystems, ranging from altered individual behaviours in native species through to alterations in ecosystem functioning (Cucherousset and Olden, 2011). Understanding the disruptions in processes, such predation and competition, caused by invasive species is important in determining the strength and direction of their impacts on native biodiversity (Tran et al., 2015; Britton, 2023) Consequently, an important first step in understanding the extent of the disruption to these processes is determining the extent of the trophic interactions of the invader with native species (Britton, 2018).

Following the introduction of an alien species into a novel ecosystem, theory suggests that if there are some prey resources that are largely unexploited then their exploitation by the introduced species will avoid inter-specific competitive interactions, increasing their probability of establishment and invasion (Juncos et al., 2015). Conversely, an introduction into an environment where the prey resources are fully exploited and/or are limited could result in competitive interactions increasing, driving a constriction in the size of the trophic niches of all species as they develop dietary specialisms (Tran et al., 2015; Jackson et al., 2016). However, this scenario can also result in each species having larger and convergent trophic niches as all species become more generalized in their resource use to maintain their energetic intake (Svanbäck and Bolnick, 2007). Where the strength of inter-specific competition is high and the alien species establishes an abundant population, some native species may then be displaced from their original trophic niche via competitive exclusion (Bøhn et al., 2008). Whilst testing these trophic niche theories for alien freshwater fishes can be challenging due to an absence of information on trophic niche sizes and positions in the pre-invasion period, this can be at least partially overcome where there are control sites available nearby with similar native fish communities present that remain uninvaded (e.g. Britton et al., 2010).

The detrimental impacts of alien species can be elevated when the impacted species are endemic, as their extinction risk rises (De Santis et al., 2021; Trombley et al., 2021). An example of such a species is the Aegean chub, Squalius fellowesii, an endemic fish of the Aegean drainages of Anatolia, Turkey, where there are concerns over the consequences of the invasion of the North American centrarchid pumpkinseed Lepomis gibbosus (Top et al., 2016a). The Aegean chub is an important socio-economic species due to their consumption by local people, with some concerns on over-exploitation (Balık et al., 2004; Tarkan et al., 2015). Other threats to their populations are habitat loss and degradation (e.g. pollution, channelization, damming), drought and excessive water extraction (Fricke et al., 2007).

The pumpkinseed is a small-bodied, warm-water fish that has been introduced firstly to Europe as an ornamental fish in the 19th century from North America (Copp and Fox, 2007) and is one of the most introduced aquatic species worldwide (Garcia-Berthou, 2007). Its high establishment success is associated with its tolerance to harsh environmental conditions, including high temperatures and hypoxia (Farwell et al., 2007). Studies on invasive pumpkinseed have been conducted on both pond (e.g. Copp et al., 2002; Villeneuve et al., 2005; Fox et al., 2011) and stream dwelling populations (Cucherousset et al., 2009; Fobert et al., 2013; Almeida et al., 2014; Top et al., 2016b; Tarkan et al., 2021). Despite pumpkinseed spreading in Turkey since the 1980s (Tarkan et al., 2015), investigations of stream-dwelling Turkish populations have focused on their habitat use, which suggested minimal evidence that their presence was impacting endemic species (Top et al., 2016b), although high plasticity in pumpkinseed life history traits was noted (Tarkan et al., 2021). There is, however, no information on pumpkinseed-endemic fish trophic interactions in Turkey. Trophic studies on stream-dwelling pumpkinseed from elsewhere in Europe have suggested the species can be highly aggressive in their interactions with native species, such as in Iberia (Almeida et al., 2014), but with their interactions with some species being minimal, such as with brown trout Salmo trutta in streams in Southern England (Copp et al., 2017).

The aim here was to thus assess the trophic interactions and consequences of invasive pumpkinseed with endemic Aegean chub (‘chub’ hereafter) by using isotopic niche as a proxy in a small Mediterranean stream in Turkey. In this stream, the chub is widely distributed and tends to be the most abundant species present, with other native and alien species having very patchy distributions and very low abundance. Where pumpkinseed is present, they also tend to be highly abundant. The trophic interactions, condition and growth rates of chub and pumpkinseed were assessed through samples collected on the stream to represent a design of a natural treatment (sites with co-habiting chub and pumpkinseed) and a control (sites with chub without pumpkinseed and with pumpkinseed without chub). Objectives were to assess the trophic ecology and niche breadth, and growth and condition of chub in pumpkinseed presence and absence, and of pumpkinseed where present, and to discuss the conservation implications.

2 Materials and methods

2.1 Study area and sites

The study stream was the Sarıçay Stream in Muğla-Milas region, south-west Anatolia (Tab. 1; Fig. 1). The stream is approximately 70 km in length and generally comprises of a range of habitat types (e.g. pools and riffles, consisting sand, gravel, mud, and rock as substrate types). Maximum depths rarely exceed 2 m and mean width of the stream is approximately 10 m. A number of alien fishes (pumpkinseed, gibel carp Carassius gibelio, common carp Cyprinus carpio, eastern mosquitofish Gambusia holbrooki, topmouth gudgeon Pseudorasbora parva) were present, along with patchily distributed native/endemic fishes (Smyrna chub Petroleuciscus ninae, loach Oxynoemacheilus sp.), plus the most abundant Aegean chub (Top et al., 2016b). Pumpkinseed was accidentally introduced into the catchment via the Geyik reservoir (GE, Fig. 1) through contaminated stockings of common carp for fisheries enhancement in the early 1990s (Top et al., 2016b). The stream is anthropogenically affected throughout by residential areas (i.e. villages), small-scale factories and sand quarries.

Table 1

Location, coordinates (Lat., Lon.), sampled species, altitude (Alt., m), temperature (T, °C), pH, dissolved oxygen (DO, mg−1 and %), conductivity (EC, μS cm−1) and salinity (S, ‰) of sampling points measured in April 2017 along the Sarıçay stream.

thumbnail Fig. 1

Location of the sampling sites in the Sarıçay stream.

2.2 Sampling design and laboratory processing

Sampling was conducted in compliance with ARRIVE guidelines (Percie du Sert et al., 2020) and all methods were carried out in accordance with the relevant guidelines and regulations, with ethical approval also granted by the Animal Experiments Local Ethics Committee of Ankara University (Decision No: 2015-14-157). There were seven field sites used, all of 100 m length (numbered from site 1–upstream to site 7–downstream along the stream) (Fig. 1; Tab. 1). Notably, sampling in reservoirs were conducted in river-type habitats in the reservoirs (i.e. in the stream reaches of the reservoirs). The physical modifications of the basin, which include two large reservoirs and several small ponds formed by disused sandpits, prevented pumpkinseed from occurring throughout the entire stream course. However, they did establish populations in the reservoirs and artificial ponds where chub could not survive, allowing us to observe different species interaction combinations, such as chub-pumpkinseed in both allopatry and sympatry. Accordingly, the study consisted of three treatments; chub in allopatry (Site 1, 4 and 5; n = 50), pumpkinseed in allopatry (Sites 2 and 3; n = 40) and both species in sympatry (Sites 6 and 7, n = 20+20) (Tab. 2). Sampling was conducted in April 2017, with fish captured using a portable electrofishing device without stop nets. In all sites, fishing was continuous within 10-min blocks to enable use of a catch per unit effort (CPUE) metric, which were used as a measure of relative fish abundance, expressed as the number of fish captured per 10 min fishing (Copp et al., 2005). Following capture, fish were euthanized via anaesthetic overdose (2-phenoxyethanol) and transported to the laboratory on an ice water slurry. Water temperature, conductivity, dissolved oxygen, salinity and pH were measured at each sampling site just prior to the fish sampling using a YSI 556 MPS probe. There were only minor differences in most environmental variables across the sampling locations (Tab. 1). The exception was in salinity and conductivity, which generally increased with distance downstream, plus low dissolved oxygen values were evident at both sites 4 and 7, due to local issues (Tab. 1). Also, samples of fish putative prey resources (submerged macrophytes, zooplankton, macrobenthos) were collected from each sampling site. Macrophytes and macrobenthos were collected with a scoop and a grab from the stream bottom and surface, while zooplankton was collected with a zooplankton net.

In the laboratory, the collected fishes were identified as either chub or pumpkinseed, measured (total length and weight, nearest mm and g, respectively). For ageing, scale samples of each species were taken from the area between lateral line and dorsal fin. All specimens were stored at −20 °C, before defrosting and a sample of dorsal muscle tissue excised from each individual for bulk stable isotope analysis (SIA). As white muscle tissue has lower variability in nitrogen isotopic signature compared to other tissues it does not require acidification to remove inorganic carbonates (Pinnegar and Polunin, 1999). The muscle samples, plus those of the fish putative prey, were then dried in an oven for 24 h at a constant temperature (60 °C) before being ground to fine powder using an agate pestle and mortar. A 1 mg sample of homogenized tissue from each fish and invertebrate material and 2 mg of plant material, each was weighed accurately (i.e. 0.150 ± 0.001) and put into 5 × 9 mm tin cups using an ultra-microbalance (Sartorius MSA3.6P-000- DM Cubis Micro Balance). Samples were analysed by an elemental analyser coupled with a continuous flow isotope ratio mass spectrometer at Akdeniz University Food Safety and Agricultural Research Centre. All stable isotope values are reported in the δ notation: δ13C or δ15N = [((Rsample/Rstandard) − 1) × 1000], where R is 13C/12C or 15N/14N ratios, and the standards are atmospheric nitrogen and PeeDee Belemnite. The secondary standards for stable N isotope USGS-34 Potassium Nitrate and for stable C isotope CH-6 sucrose and total C and N ratio were estimated by using aspartic acid, cyclohexanone-2,4-DNPH, acetanilide and nicotinamide standards. The standard errors of the replicates of all analyses were 0.03‰ for δ13C and 0.11‰ for δ15N. The ageing of the collected scales was through counting true annuli (Vilizzi et al., 2013). Scale impressions of 10 scales for each specimen on acetate strips were read on a micro-projector (magnification: ×48). Two independent readers estimated the ages without knowing prior biological status of the fish and a final evaluation was made when age estimates differed. If the disagreement continued, the sample was excluded from the dataset. Rejected samples together with regenerated scales, meant that 3.4% of all sampled fish were removed from further age-based analyses. After age estimation, total scale radius and the distances from the scale focus to each annulus of one scale per fish was measured for the back-calculation of length at age (Bagenal and Tesch, 1978).

Table 2

Mean TL (mm), TW (g), δ13C, δ15N, TP and δ13Ccorr (± 95% CI) of chub and pumpkinseed by treatment (allopatry and sympatry) in the sampling sites in Sarıçay stream.

2.3 Data analyses

Somatic growth rates of both fishes were calculated using the Fraser-Lee equation (Francis, 1990):

where Lt is TL when annulus t was formed, TLc is TL at capture, St is the distance from scale focus to the annulus t, R is scale radius, and c is the intercept on the length axis from the linear relationship of TL versus scale radius. A fixed body-scale intercept (c = 17.7 mm for chub 22.7 mm for pumpkinseed) was used for each sampled population to reduce bias due to differences in the size distribution. The relative growth of both species across the sampling locations were compared through a relative growth index (GI) following (Hickley and Dexter, 1979). This involved use of the Walford (Walford, 1946) method to obtain a straight line by plotting TL at mean age n of all specimens against TL at age n + 1, with GI then calculated as the mean value of the growth in each age class of the species for a certain site relative to TLs at age, using the following formulae:

where L∞ = lt/(1–k) and lt is the intercept, ln is the TL at age n and k is the slope of the Walford plot (Hickley and Dexter, 1979) and:

where TLoi and TLri × 100 are the observed and reference mean TLs respectively of the species at age i. Because of observed high levels of variability in the estimated length at older ages in the reviewed dataset, which suggested some potential issues of ageing accuracy and precision (Beamish and McFarlane, 1983), only comparable ages (i.e. the first four age classes) of the fish from each site were used in calculations. The relative body condition (RC) of Le Cren (Le Cren, 1951) was used to measure variation in body condition using according to W/W′, where W is the actual weight of an individual and W′ is the expected weight from the WL relationship. RC values >1 or <1 show that specimens are in better or worse condition than the other fish at the same TL range. For the proper use of this index, all populations are required to be sampled at the same time of year (Knaepkens et al., 2002) hence, samples were collected from all sampling sites on the same day.

For the fish stable isotope data, with the δ13C values were arithmetically transformed prior to analysis to account for muscle lipid content (Kiljunen et al., 2006). As the fish putative prey SI data revealed significant differences in δ13C between the sites (F = 16.90, df = 13, P < 0.01) (Supplementary material, Table S1) then, for each site, the fish δ13C data were standardised by their conversion to δ13Ccorr according to Olsson et al. (2009):

where δ13Ccorr is the corrected carbon isotope ratio of the individual fish, δ13Ci is the uncorrected isotope ratio of that fish, δ13Cmeaninv is the mean invertebrate isotope ratio and CRinv is the carbon range of the invertebrate (δ13Cmax − δ13Cmin) (Olsson et al., 2009). The δ15N were then converted to trophic position (TP) to provide more meaningful ecological metric:

where TPi is the trophic position of the individual fish, δ15Ni is the isotopic ratio of that fish, δ15Nbase is the isotopic ratio of primary consumers (i.e. baseline invertebrates), 3.4 is the fractionation between trophic levels and 2 is the trophic position of the baseline organism (Post, 2002).

To then test the corrected SI data, the site-level data were grouped into ‘experimental’, and ‘control’ stretches, where the experimental group were sites where chub co-habited with pumpkinseed, and the control groups were sites where pumpkinseed and chub were in allopatry. Standard ellipse areas (SEAs) were calculated for each group using the SIBER package in R (Jackson et al., 2011). SEA is a bivariate estimate of the core isotopic niche width considering 40% of the central data points (Jackson et al., 2011, 2012, 2016). Bayesian estimates of SEA (SEAb) were calculated using a Markov chain Monte Carlo simulation (104 iterations per category) to test for differences in sizes of isotopic niches by calculating the probability that the relative posterior distributions of the niche sizes were significantly smaller or larger between the species (α = 0.05) (Jackson et al., 2011, 2012). Thus, no overlap in the 95% confidence intervals of SEAb indicated a significantly smaller or larger niche size (depending on the direction of the difference between the species). Isotopic niche overlaps (%) between the species were calculated using the SEAc in SIBER (subscript ‘c’ indicates a small sample size correction was used; Jackson et al., 2012). The extent of niche overlap between the species used pumpkinseed as the comparator species.

A permutational univariate analysis of variance (PERANOVA) was used to test the significance of differences of CPUE, standard length, weight, TP and δ13Ccorr between treatments using PERMANOVA 1.0.1. + add-in to PRIMER version 6.1.11 (PRIMER-E Ltd., Plymouth, UK, Anderson et al., 2008). This procedure was performed with Type III sums of squares following normalization of the data and was based on a Euclidian distance matrix and 9999 permutations of the residuals under a reduced model. The use of PERMANOVA versus traditional parametric analysis of variance (ANOVA) is advantageous as the stringent assumptions of normality and homoscedasticity in the data are substantially relaxed in PERMANOVA, enabling its use with real-world ecological data sets.

3 Results

3.1 Fish sample characteristics

There were 312 chub sampled from the five sites where they were present, varying from 31 individuals at site 7 (pumpkinseed present) to 87 at site 4 (pumpkinseed absence), with fish present between 75 and 216 mm in length and 4.6 to 125.0 g in mass (Tab. 2). Pumpkinseed were collected from 4 sites, with sample sizes varying from 82 at site 7 (chub present) to 406 at site 2 (chub absent); all pumpkinseed were between 23 and 124 mm in length and 0.2–39.2 g in mass (Tab. 2).

Both mean TL and TW of chub were significantly higher in pumpkinseed absence than presence (TL: F = 3.24, df = 1, P < 0.02; TW: F = 2.81, df = 1, P < 0.01). This was true for pumpkinseed, where both mean TL and TW were significantly higher in the absence of chub (TL: F = 19.61, df = 1, P < 0.001; TW: F = 26.22, df = 1, P < 0.001) (Fig. 2). There were significant differences in relative abundance (as CPUE) for both species, being significantly higher in chub where pumpkinseed were absent compared with where pumpkinseed were present (F = 6.50, df = 1, P < 0.01). Pumpkinseed were also more significantly abundant in the absence of chub (F = 22.36, df = 1, P < 0.001) (Fig. 2). The relative condition was significantly higher for chub in pumpkinseed absence (F = 3.28, df = 1, P < 0.01) and for pumpkinseed in chub absence (F = 15.40, df = 1, P < 0.0001 (Fig. 2). Growth rates (as GI) of both species were higher when sympatry (both species together) (chub: F = 49.10, df = 1, P < 0.01; pumpkinseed: F = 21.14, df = 1, P < 0.001) (Fig. 2).

thumbnail Fig. 2

Catch per unit effort (CPUE), relative condition (RC) and growth index (GI) of Aegean chub and pumpkinseed in the sampling sites.

3.2 Stable isotope ecology and metrics

Comparison between treatment groups revealed chub TP and δ13Ccorr was significantly higher in pumpkinseed presence than absence (TP: F = 5.323, df = 1, P < 0.01; δ13Ccorr: F = 24.876, df = 1, P < 0.001) (Tab. 2). Pumpkinseed TP was also highest in chub presence (F = 25.431, df = 1, P < 0.001) (Tab. 2). In general, the relationship between TL and both TP and δ13Ccorr were not significant, with the exception of positive relationship for δ13Ccorr in pumpkinseed (Fig. S1).

The chub had similar sized isotopic niches (probability 0.21) between pumpkinseed presence and absence, and with some shift in niche position between the contexts (niche overlap was 45% between pumpkinseed presence/absence) (Tab. 3; Fig. 3). Pumpkinseed in chub presence had a larger isotopic niche (probability 0.23) than in chub absence, and overlapped by 67% (Tab. 3, Fig. 3).

Table 3

Isotopic niche sizes (as standard ellipse areas, SEAc and SEAb, with 95 % CI) and ellipses overlap for intra and inter-specific status.

thumbnail Fig. 3

Isotopic niches (as standard ellipse areas, SEAc) for (a) chub allopatry vs. pumpkinseed allopatry, (b) chub sympatry vs. pumpkinseed sympatry, (c) chub in allopatry vs. sympatry, (d) pumpkinseed allopatry vs. sympatry. TP = trophic position (calculated from δ15N data), δ13Ccorr = corrected value of δ13C.

4 Discussion

Comparison of the isotopic niche sizes of the two fishes across the different contexts suggested that the consequences of pumpkinseed presence for chub were minor, with similar sized niches in their presence/absence and maintaining some of their isotopic space between pumpkinseed presence/absence (45%). Despite chub sharing their isotopic niches with pumpkinseed where they are co-habited, there were few apparent consequences for their growth or body condition.

Although empirical evidence showed that non-native fishes can have detrimental ecological impact for native fishes and communities as a result of trophic interactions (Cucherousset and Olden, 2011; Britton, 2023), co-occurrence with non-native pumpkinseed in the present study revealed no significant impact on chub growth and relative condition. In fact, a significant increase in growth but reduced condition was observed in both species in case of co-occurrence. This was supported by the data from a previous study of habitat use of both species in one of the same sampling sites (site 6) in Sarıçay stream that showed a microhabitat partition (Top et al., 2016b). These outputs were also consistent with Stakėnas et al. (2013) and Jackson et al. (2016) who, despite showing higher habitat and resource partitioning between brown trout and pumpkinseed in a similar stream ecosystem in England, also found little evidence of any ecological impacts arising from these processes. The results are, however, contrary to Copp et al. (2017) who used also stable isotope analysis in experimental pond conditions, detecting high niche overlap and reduced niche size in native gudgeon Gobio gobio and roach Rutilus rutilus in pumpkinseed presence. In a similar experimental study on impact of pumpkinseed on perch Perca fluviatilis, no effect on growth was observed in either species, but both species shifted their diet to avoid competition (Fobert et al., 2011). These results support an earlier study in the stream on habitat interactions of pumpkinseed with endemic fishes suggesting little evidence for impact (Top et al., 2016b), meaning there is currently little need for further management actions to mitigate the effects of their ecological interactions.

The stable isotope analysis approach used in this study complements previous trophic studies on this stream that assessed the trophic interactions between non-native topmouth gudgeon, pumpkinseed and native/endemic species using traditional stomach analyses (Karakuş, 2014). Then, significant feeding overlaps were apparent between the two invaders, and between pumpkinseed and native Oxynoemacheilus sp. (Karakuş, 2014). However, since this study, the topmouth gudgeon population has declined markedly and Oxynoemacheilus sp. was captured in very low abundance, meaning it was not possible to test their isotopic niches in the present study. Whilst other studies have noted that dietary shifts in native fishes might occur after an introduction of non-native species e.g. Britton et al. (2010), such changes were not clearly apparent in the present study. This might be related to biotic resistance of the stream, evidenced by minimal habitat/food interactions of pumpkinseed with native species or their presence has not been sufficient to exert a long-term shift in stream community structure. Also, the data have the caveat that they were non-replicated field samples that were only sampled once and subject to uncontrolled environmental conditions (e.g., fish abundance, food availability, ecosystem size).

Pumpkinseed are highly invasive in much of Europe, especially in the south e.g. Iberian Peninsula, (Godinho et al., 1997) where they mainly invade impounded reaches of large river systems (Mesquita et al., 2006; Almeida et al., 2009). Though not so wide-spread, pumpkinseed has established populations in Turkey (Tarkan et al., 2015) but still there is no report on their invasive populations causing ecological impacts, which could also be due to lack of relevant studies. Future climate change models predicted a country-scale range expansion of pumpkinseed and potential niche overlaps with some common endemic cyprinid species (Emiroğlu et al., 2023), with likely increases in dispersal via fluvial networks. Thus, the negligible ecological impact found in the present study should be approached with caution, as pumpkinseed is likely to spread to other systems under future scenarios of climate change where detrimental impacts could be incurred. However, this may just be suggestive of future impacts, as climate change would also affect the distribution, abundance and growth of native species (e.g. Van der Putten, 2010) and impact of non-native species could be context-dependent (i.e. based on the recipient ecosystem) (Catford et al., 2021).

Detrimental impacts of non-native species on native species have been widely recognized and reported (Gozlan et al., 2010; Britton, 2023), mainly through processes such as increased predation and interspecific competition, but also through habitat modifications and interactions resulting in interference (Gozlan et al., 2010; Cucherousset and Olden, 2011; Almeida et al., 2014). In the literature, there is usually a bias to focus on the cases where high abundances of invaders occur, but this inherently results in high ecological impacts, which potentially overstates the ecological consequences of non-native species (Gozlan, 2008; Jackson et al., 2015). It is therefore highly crucial for prioritizing proper management actions (Britton et al., 2023) to report such cases where impact of non-native fishes is little/relatively benign (e.g. Fobert et al., 2011) or biotic resistance of the native species and ecosystems is considerably high. The example provided here of the pumpkinseed in the Sarıçay Stream is such an example, where their impacts on the endemic Aegean chub appeared to be minimal, despite their relatively high resource sharing in sites where they co-occur.

Supplementary Material

Table S1. Mean, min, max δ13C, δ15N of macroinvertebrates in the sampling sites in Sarçéay stream.

Fig. S1. Relationships of TP and d13Ccorr versus total length (TL) of Aegean chub (left side) and pumpkinseed (right side).

Access here

Acknowledgements

We thank Phillip J. Haubrock for revision of figures. This study was supported by both The Scientific & Technological Research Council of Turkey (TÜBİTAK) (Project No: 115Y660) and Scientific Research Project Fund of Muğla Sıtkı Koçman University under the project number 16/153.

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Cite this article as: Karakuş U, Tarkan AS, Yalçın-Özdilek Ş, Top-Karakuş N, Partal N, Britton JR. 2023. Interactions of invasive pumpkinseed with a riverine endemic fish indicate trophic overlap but minimal consequences for somatic growth and condition. Knowl. Manag. Aquat. Ecosyst., 424, 13.

All Tables

Table 1

Location, coordinates (Lat., Lon.), sampled species, altitude (Alt., m), temperature (T, °C), pH, dissolved oxygen (DO, mg−1 and %), conductivity (EC, μS cm−1) and salinity (S, ‰) of sampling points measured in April 2017 along the Sarıçay stream.

In the text
Table 2

Mean TL (mm), TW (g), δ13C, δ15N, TP and δ13Ccorr (± 95% CI) of chub and pumpkinseed by treatment (allopatry and sympatry) in the sampling sites in Sarıçay stream.

In the text
Table 3

Isotopic niche sizes (as standard ellipse areas, SEAc and SEAb, with 95 % CI) and ellipses overlap for intra and inter-specific status.

In the text

All Figures

thumbnail Fig. 1

Location of the sampling sites in the Sarıçay stream.
In the text
thumbnail Fig. 2

Catch per unit effort (CPUE), relative condition (RC) and growth index (GI) of Aegean chub and pumpkinseed in the sampling sites.
In the text
thumbnail Fig. 3

Isotopic niches (as standard ellipse areas, SEAc) for (a) chub allopatry vs. pumpkinseed allopatry, (b) chub sympatry vs. pumpkinseed sympatry, (c) chub in allopatry vs. sympatry, (d) pumpkinseed allopatry vs. sympatry. TP = trophic position (calculated from δ15N data), δ13Ccorr = corrected value of δ13C.
In the text

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