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All issues No 425 (2024) Knowl. Manag. Aquat. Ecosyst., 425 (2024) 8 Full HTML
Issue
Knowl. Manag. Aquat. Ecosyst.
Number 425, 2024
Freshwater ecosystems management strategies
Article Number 8
Number of page(s) 11
DOI https://doi.org/10.1051/kmae/2024004
Published online 12 April 2024

© M. Gimenez and J. Cucherousset, Published by EDP Sciences 2024

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

Stocking programs are primarily designed to enhance fish stocks by releasing hatchery-reared or wild-borne individuals (Lorenzen et al., 2012; Sass et al., 2017) to optimize the angling experience (Cowx, 1994). In addition, to satisfy catch-challenge motivations of anglers, they often involve the release of fish species, especially non-native predators, which represent a fishing challenge, but are also known to affect recipient communities and ecosystems (Eby et al., 2006; Rudman et al., 2016). In fact, the stocking of predatory fish can result in a trophic cascade due to increased top-down control (Eby et al., 2006), and lead to altered trophic linkages through changes in prey community structure (Tiberti et al., 2014) that can subsequently affect ecosystem functions such as nutrient cycling and primary production (Schindler et al., 2001). Furthermore, individuals used in stocking programs often come from fish farms and are therefore reared in conditions that are very different from the ecosystems where they are released. These differences have multiple consequences on stocked fish (Milla et al., 2021), including change in foraging behavior of predatory fish fed with artificial diets in hatchery (Huntingford, 2004; Milla et al., 2021). Thus, stocking of predatory fish species, and especially hatchery-reared individuals, can be a possible threat to the long-term sustainability of freshwater biodiversity and ecosystems. However, stocking can also contribute to the recovery of fish stocks and provide benefits for fisheries (Lorenzen et al., 2012; Amoroso et al., 2017; Johnston et al., 2018), but this depends on the ecological characteristics of the recipient ecosystem, which is rarely considered (Claussen and Philipp, 2023).

Habitat enhancement measures that aim to improve fish stocks have received less consideration by freshwater managers than stocking (Lorenzen, 2014; Schindler and Hilborn, 2015; Sass et al., 2017). Nearshore habitats are often impacted by recreational activities or intentionally removed by fishery managers (Jennings et al., 2003), although it can affect the growth and cause a decline of fish populations (Sass et al., 2006). Increased investment in habitat enhancement and protection could have beneficial effects for the ecosystems, by increasing its carrying capacity, and for the fisheries, by improving the productivity of exploited fish stocks (Radinger et al., 2023) through increased natural recruitment in some popular fish species (Roni et al., 2008; Nilsson et al., 2014). In fact, structured habitat are particularly important for freshwater fishes because they can be used for foraging and predator avoidance (Sass et al., 2023). Compared to stocking, habitat enhancement appears to be a more integrative and sustainable management, because adding habitats can directly support biodiversity and contribute by extension to the ecosystem functioning and services (Dobson et al., 2006; Harrison et al., 2014; Radinger et al., 2023). In lake, littoral habitats such as coarse woody habitat (CWH), including for example logs, root wads and stumps, provide key habitats not only for fish, but also for invertebrates or periphyton, and affect a wide range of ecological functions (Harmon et al., 2004), such as ecosystem productivity and nutrient cycling (Czarnecka, 2016). Furthermore, the addition of this type of physical habitat structure represents one of the most common methods used to restore freshwater habitats (Cooke et al., 2023), and it is especially beneficial in ecosystem lacking habitats (Sass et al., 2023). The deployment of physical habitat structures is not limited to the addition of woody habitats, and many structures made from artificial materials have also been used, particularly in recreational fisheries (Bolding et al., 2004). Such structure are also subjected to debate because of a potential release of contaminants (e.g., with plastic habitats Cooke et al., 2023). Thus, enhancing the quality and quantity of littoral habitats might limit some of the negative effects of predatory fish stocking by providing refuge to prey community and maintaining some important ecosystem functions, but this remains to be tested.

The main objective of the present study was to determine whether habitat enhancement can limit the ecological impacts of predatory fish stocking. To study the effect of stocking, we used an experimental mesocosm approach and hatchery-reared juvenile largemouth bass Micropterus nigricans, a predator and a popular game fish for anglers (Donaldson et al., 2011), as a model species. We assessed the ecological consequences of two different habitat enhancements, a habitat made of natural materials and a habitat made of natural and artificial materials (i.e., CWH and CWH within an iron cage), prior and after fish stocking. First, we predicted that stocking will have a strong impact on the composition of prey communities through predation, and consequently modify ecosystem functioning. Then, we predicted that habitat enhancement will limit the ecological impacts of stocking by supporting prey communities (i.e., refuge from predation and additional physical habitat), thereby maintaining some important ecosystem functions. Finally, we have also predicted that the effect of habitat enhancement will differ depending on the materials used and the complexity of the artificial structure.

2 Materials and methods

2.1 Model system

We used the stocking of non-native largemouth bass in gravel pit lakes in southwestern France as a case study for this experiment. In many European countries, gravel pit lakes are becoming more and more common in the landscape and represent valuable ecosystems for recreational fisheries (Matern et al., 2019; Meyerhoff et al., 2019), but the availability of structured littoral habitats for fish is often limited. In the study area, stocking is an important driver of fish community structure in gravel pit lakes (Zhao et al., 2016; Gimenez et al., 2023; Garcia et al., 2023). Largemouth bass is one of the most notorious and highly introduced species in freshwater ecosystems (Brown et al., 2009; Donaldson et al., 2011) and its introduction can affect recipient community through trophic cascade (Drenner et al., 2002; Jackson, 2002). Largemouth bass is well distributed in the study area, targeted by many recreational anglers and local angling authorities have recently implemented a regulation whereby catch-and-release angling is mandatory for this non-native species. The littoral habitat is particularly important for the all life stages of the species (Brown et al., 2009), but gravel pit lakes are usually characterized by steep slopes and higher littoral depth with a lack of macrophytes and deadwood compared to natural lakes (Emmrich et al., 2014). Thus, littoral habitat enhancement could be particularly efficient in such ecosystems (Radinger et al., 2023), and this is particularly true for this species.

2.2 Experimental design

We used an experimental approach (24 outdoors mesocosms, circular cattle tanks, 550 L) based on four treatments: i) nor fish stocking or habitat enhancement (hereafter “no stocking”), ii) fish stocking and no habitat enhancement (“stocking only”), iii) fish stocking and habitat enhancement with a coarse woody habitat (CWH) (“stocking and CWH”) and iv) fish stocking and habitat enhancement with a caged-CWH (“stocking and caged-CWH”). Each treatment was replicated six times and we used a block design with six blocks with each block containing the four treatments (Závorka et al., 2020) (Fig. S1). CWH habitat enhancement consisted in a brush pile consisting of 2 kg of deadwood (thin branches < 1.5 cm diameter of downy oak Quercus pubescens) grouped into circular bundles (40 × 20 × 20 cm) (Fig. S2a). Caged-CWH habitat enhancement consisted in the same bundle of deadwood that was placed within an iron cage (40 × 40 × 30 cm, Biohut®, ECOCEAN, Montpellier, France) (see details of the structure in Fig. S2b). Specifically, the cage was made of iron coated with a zinc-aluminum alloy. This device is deployed to restore nurseries in marine ecosystems and was used here to increase habitat complexity and verticality compared to CWH.

On April 29, 2021, 5 cm of gravel and 500 L of unfiltered water from a nearby gravel pit lake were added to each mesocosm. On June 10, 2021, each mesocosm was inoculated with phytoplankton and zooplankton collected from three gravel pit lakes (Téoula, Soulance and Lamartine lakes) located in the floodplain of the Garonne river (Haute-Garonne, France) (Zhao et al., 2016). Phytoplankton and zooplankton inoculum were collected using nets with a mesh size of 10 and 50 µm, respectively. From May 18, 2021 to July 15, 2021, several additions of multiple macroinvertebrate taxa were made. In each mesocosm, 12 pond snails (immature Lymnaea stagnalis) reared in aquarium and two mesh bags containing 5 g of leaves of black poplar Populus nigra that have been placed in a nearby gravel pit lake for 21 days were added to each mesocosm. On June 28 and July 15, 2021, two additional inoculums of macroinvertebrates collected in the Ariège river using a dip net (500 µm mesh size) were added.

The experiment was composed of two periods. The first period (before stocking) started on November 22, 2021, when habitat enhancements were installed and lasted 34 weeks. The second period (after stocking) started on July 20, 2022 with fish stocking and ended on September 6, 2022 (duration: eight weeks). From July 4, 2022 until the end of the experiment, nets (1.5 mm mesh size) were placed on the mesocosms to prevent fish escapees and potential predations. Summer 2022 was particularly warm and dry with multiple heat waves. Therefore, water was added to all mesocosms during the warmest period to compensate evaporation and an air bubbling system was placed in each mesocosm to avoid hypoxia.

2.3 Fish stocking

For all mesocosms with stocking, five hatchery-reared young-of-the-year (YOY) largemouth bass were introduced (ntotal = 90 individuals) on July 20, 2022. In the hatchery, they were fed daily with pelletized food. To avoid handling stress, the total mass of largemouth bass introduced in each mesocosm was measured as a batch (mean = 33.6 g ± 2.4 SD) and did not differ significantly between treatments (LMM, χ2(2,17) = 0.11, p = 0.95, Fig. S3).

2.4 Mesocosm monitoring and sampling

On September 6, 2022, fish were collected from each mesocosm and subsequently euthanized individually using an overdose of anesthetic (benzocaine, 25 mg.L−1). The number of fish in each mesocosm was counted and individual size (nearest mm) and mass (nearest 0.01 g) were measured. Then, the survival rate was calculated based on the number of fish introduced and the number of fish found alive at the end of the experiment. We also calculated the scaled-mass index as an index of body condition following Peig and Green (2009):

where Wi and Li are the weight and length of each fish, respectively. In our case, L0 was the arithmetic mean of the length of all fish measured. To compute the bSMA, we applied an SMA regression (using the ‘lmodel2’ function in R) to log-transformed weight and length values to determine the slope of the fitted line.

Regarding prey communities, zooplankton was sampled by filtering 20 and 40 L of water using a 80 µm mesh sieve, respectively for each period (i.e., before stocking on July 19, 2022, and after stocking on September 5, 2022). Samples were stored in 96% ethanol and subsequently filtered again through a 160 µm mesh sieve prior to counting. Zooplankton was identified to the lowest taxonomic level using a binocular magnifier (Leica MZ7.5 StereoZoom Microscope). A total of eight taxa (Chaoboridae, Cyclopoïda, Calanoïda, Ceriodaphnia, Simocephalus, Daphnia, Bosminidae, Chydoridae and Nauplii) were identified and counted. We then calculated zooplankton richness (number of taxa) and density (ind.L−1). Benthic macroinvertebrates were sampled using three plastic trays buried in the substrate installed in the mesocosms on June 10 and July 22, 2022, respectively. Macroinvertebrates were collected using a sieve (before stocking on July 19, 2022, and after stocking on September 5, 2022) and stored in 96% ethanol before laboratory analysis. Macroinvertebrates inhabiting CWH and caged-CWH were also collected, but only at the end of the experiment (September 6, 2022) to avoid overly destructive sampling before fish introduction. Macroinvertebrates were then identified to the lowest taxonomic level (mainly Family) using a binocular magnifier (Leica MZ7.5 StereoZoom Microscope). A total of 16 macroinvertebrate taxa (Lymnaea, Physa, Ancylidae, Asellidae, Gammaridae, Planarian, Chironomidae, Culicidae, Dytiscidae, Hydrachnidia, Annelida, Ceratopogonidae, Baetidae, Corixidae, Corbicula, Mesovellidae) were identified and counted. We then calculated macroinvertebrate richness (number of taxa) and density (ind.mesocosm−1).

Regarding ecosystem functioning, parameters were measured the day before community sampling to limit potential perturbations. Pelagic primary production was assessed by measuring total chlorophyll-a concentration (µg chl-a.L−1) using a portable fluorometer (AlgaeTorch; BBE moldaenke GmbH) before fish stocking. After stocking, chlorophyll-a concentration (µg chl-a.L−1) was quantified by filtering water samples (250 mL) due to high concentration in some mesocosms. Benthic primary production was assessed by measuring chlorophyll-a concentration (µg chl-a.cm−2) on three ceramic tiles (20 × 10 cm) using a portable fluorometer (BenthoTorch; BBE moldaenke GmbH; Kahlert and McKie, 2014). The decomposition rate of leaf litter was quantified by measuring breakdown of 3 g (± 0.02 g) of air-dried black poplar leaves (details in Alp et al., 2016; Závorka et al., 2020) placed into mesocosms. Leaf decomposition corresponded to the decomposition rate for each tank, which was calculated as follows (Lecerf et al., 2005):

where Mf is final and Mi is initial oven-dried mass of leaf litter, T the duration of leaf exposure in mesocosms (39 and 45 days for each period, respectively).

2.5 Statistical analyses

Overall, a similar general approach was used to test the effects of treatment on each response variables. We used mixed effects models (LMM and GLMM depending on data distribution, see details in Tab. S1) using treatment as a fixed variable and experimental block as a random variable. The response variables used in the models were those measured directly during the sampling campaign and community identification, except for community structures, where non-metric multidimensional scaling (nMDS with untransformed densities of each taxa and Bray-Curtis index; Kruskal, 1964) were performed and the first two axes used as response variables (Tab. S1). When response variables were significant, pairwise comparisons were subsequently used to identify differences between treatments (Tukey post-hoc test). All analyses were performed with R v. 4.2.2 (R Core Team, 2022) using lme4 v. 1.1-32 (Bates et al., 2015) for LMM and GLMM, emmeans v.1.8.5 (Lenth et al., 2023) for pairwise comparisons, vegan v. 2.6-4 (Oksanen et al., 2022) for nMDS.

3 Results

3.1 Ecological effects before fish stocking

Zooplankton communities were dominated by Cyclopoïda (mean = 128.2 ind.L−1 ± 316.6 SD), Bosminidae (mean = 114.2 ind.L−1 ± 391.5 SD) and Ceriodaphnia (mean = 83.3 ind.L−1 ± 303.0 SD) (Fig. S4a). There were significant differences between treatments in the richness and density (LMM, χ2(3,23) = 10.52, p < 0.05 and χ2(3,23) = 29.78, p < 0.0001, respectively), with a significantly lower richness in stocking and caged-CWH treatment compared with stocking only treatment (Tukey post hoc test, p < 0.05, Fig. S5a). Density was significantly lower in caged-CWH treatment compared to no stocking and, stocking and CWH treatments (Tukey post hoc test, p < 0.01, Fig. S5b). There were no significant difference between treatments in zooplankton community structure (LMM, χ2(3,23) = 1.62, p = 0.65 and χ2(3,23) = 0.43, p = 0.93, for the two axes, respectively, Figs. 1a, 1c, and 1e).

Macroinvertebrate communities were largely dominated by Asellidae (mean = 3376.7 ind.mesocosm−1 ± 4765.7 SD), followed by Chironomidae (mean = 274.2 ind.mesocosm−1 ± 786.5 SD) and Annelida (mean = 250.4 ind.mesocosm−1 ± 1182.6 SD) (Fig. S6a). Richness and density differed significantly between treatments (LMM, χ2(3,23) = 12.57, p < 0.01 and χ2(3,23) = 14.14, p < 0.01, respectively), and both were significantly lower in the stocking and caged-CWH compared with no stocking and stocking only treatments (Tukey post hoc test, p < 0.05, Fig. S7 a-b). Community structure significantly differed between treatments (Fig. 1b) along the two axis (LMM, χ2(3,23) = 11.82, p < 0.01 and χ2(3,23) = 28.37, p < 0.0001, respectively for the two axes, Fig. 1d and f). Along the first axis, stocking only and stocking and CWH treatments differed significantly (Tukey post hoc test, p < 0.05, Fig. 1d), and along the second axis, stocking and caged-CWH treatment differed significantly with treatments without habitat enhancement (Tukey post hoc test, p < 0.01, Fig. 1f).

Treatments did not differ significantly for the pelagic and benthic primary production (LMM, χ2(3,23) = 5.64, p = 0.13 and χ2(3,23) = 3.54, p = 0.32, respectively, Fig. 2a-b). However, decomposition of organic matter differed significantly between treatments (LMM, χ2(3,23) = 15.37, p < 0.01), with a significant lower decomposition rate in the stocking and caged-CWH treatment compared with the no stocking treatment (Tukey post-hoc test, p < 0.01, Fig. 2c).

thumbnail Fig. 1

Effects of management practices before stocking on community structure quantified using (a-b) a non-metric multidimensional scaling (nMDS) analysis and analyzed along the (c-d) first and (e-f) second axes. Zooplankton communities are represented on the left side of the panel and macroinvertebrate communities on the right side. Different letters indicate significant difference between treatments (Tukey post hoc test, p < 0.05).
thumbnail Fig. 2

Ecological effects of treatments before stocking on (a) pelagic primary production, (b) benthic primary production and (c) decomposition rate. Different letters indicate significant difference between treatments (Tukey post-hoc test, p < 0.05).

3.2 Ecological effects after fish stocking

At the end of the experiment, there was a significant difference in scaled-mass index between treatments (GLMM, χ2(2,70) = 10.64, p < 0.01, Fig. 3a), with fish displaying significantly lower scaled-mass index in mesocosms with caged-CWH compared to stocking only (Tukey post-hoc test, p < 0.01, Fig. 3a). The survival rate did not differ significantly between treatments (LMM, χ2(2,16) = 3.56, p = 0.17, Fig. 3b).

Zooplankton communities were largely dominated by Bosminidae (mean = 147.7 ind.L−1 ± 267.6 SD), followed by Cyclopoïda (mean = 36.5 ind.L−1 ± 76.2 SD) and Ceriodaphnia (mean = 21.9 ind.L−1 ± 66.5 SD) (Fig. S4b). There were no significant difference between treatments in zooplankton richness and density (LMM, χ2(3,23) = 6.15, p = 0.10 and χ2(3,23) = 5.59, p = 0.15, respectively, Fig. S5 c-d). Community structure significantly differed between treatments (Fig. 4a) along the first axis of the nMDS, (LMM, χ2(3,23) = 22.16, p < 0.0001, Fig. 4c), with no stocking and stocking and caged-CWH treatments significantly different with stocking only treatment (Tukey post hoc test, p < 0.01 and p < 0.05, respectively, Fig. 4c). There was no significant difference between treatment along the second axis of the nMDS (LMM, χ2(3,23) = 3.19, p = 0.36, Fig. 4e).

Macroinvertebrate communities were largely dominated by Asellidae (mean = 2022.1 ind.mesocosm−1 ± 2896.5 SD), followed by Annelida (mean = 479.2 ind.mesocosm−1 ± 1382.8 SD) and Physa (mean = 52.9 ind.mesocosm−1 ± 178.5 SD) (Fig. S6b). Richness differed significantly between treatments (LMM, χ2(3,23) = 49, p < 0.0001), with a significantly lower richness in the stocking and caged-CWH treatment compared to all other treatments (Tukey post hoc test, p < 0.01, Fig. S7c). However, density did not differ significantly between treatments (LMM, χ2(3,23) = 7.20, p = 0.07, Fig. S7d). Community structure significantly differed between treatments (Fig. 4b), but only on the first axis of the nMDS (LMM, χ2(3,20) = 64.58, p < 0.0001 and χ2(3,20) = 2.40, p = 0.49, respectively for the two axes, Fig. 4d and f), with the stocking and caged-CWH treatment being significantly different from all other treatments (Tukey post hoc test, p < 0.001, Fig. 4d).

Macroinvertebrate communities in habitat were largely dominated by Physa (mean = 171.7 ind.hab−1 ± 370.1 SD) and Asellidae (mean = 59.5 ind.hab−1 ± 130.1 SD) (Fig. S8). For the two treatments with habitat enhancement action, a large proportion of the macroinvertebrates sampled were found within the habitat, despite a high variability between samples (for stocking and CWH treatment: mean = 82.4 ± 19.4% and ranging from 45.7 to 96.9%; for stocking and caged-CWH treatment: mean = 50.2 ± 47.8% and ranging from 6.0 to 100%). No significant difference in richness was found between the two habitat devices (LMM, χ2(1,11) = 1.71, p = 0.19, Fig. S9a), but density of macroinvertebrates in caged-CWH was lower compared to density in CWH (LMM, χ2(1,11) = 8.96, p < 0.01, Fig. S9b). There was no significant difference in macroinvertebrate community structure in the added habitats (Fig. 5a) along the first axis (LMM, χ2(1,11) = 0.13, p = 0.72, Fig. 5b) while a significant difference was found between CWH and caged-CWH along the second axis (LMM, χ2(1,11) = 24.92, p < 0.0001, Fig. 5c).

Regarding ecosystem functioning, pelagic primary production differed significantly between treatments (LMM, χ2(3,23) = 52.78, p < 0.0001) with higher values observed in stocking and caged-CWH than in other treatments (Tukey post-hoc test, p < 0.001, Fig. 6a). Benthic primary production did not differ significantly between treatments (LMM, χ2(3,23) = 0.59, p = 0.90, Fig. 6b). The decomposition rate of organic matter differed significantly between treatments (LMM, χ2(3,23) = 14.70, p < 0.01), with a significant lower decomposition rate in the stocking and caged-CWH treatment compared to the no stocking treatment (Tukey post-hoc test, p < 0.01, Fig. 6c).

thumbnail Fig. 3

Treatment effects on (a) scaled-mass index and (b) survival rate of juvenile largemouth bass Micropterus nigricans, at the end of the experiment. Different letters indicate significant difference between treatments (Tukey post-hoc test, p < 0.05).
thumbnail Fig. 4

Effects of management practices after stocking on community structure quantified using (a-b) a non-metric multidimensional scaling (nMDS) analysis and analyzed along the (c-d) first and (e-f) second axes. Zooplankton communities are represented on the left side of the panel and macroinvertebrate communities on the right side. Different letters indicate significant difference between treatments (Tukey post hoc test, p < 0.05).
thumbnail Fig. 5

Macroinvertebrate community structure within the added habitats after stocking quantified using (a) a non-metric multidimensional scaling (nMDS) analysis and analyzed with the (b) first and (c) second axes. Different letters indicate significant difference between treatments (ANOVA, p < 0.05).
thumbnail Fig. 6

Ecological effects of management practices after stocking on (a) pelagic primary production, (b) benthic primary production and (c) decomposition rate. Different letters indicate significant difference between treatments (Tukey post-hoc test, p < 0.05).

4 Discussion

Using an experimental approach, this study aimed to assess whether habitat enhancement, as a management practice, can help limiting the ecological impacts caused by the introduction of a non-native predatory species. First, we found that, prior to fish stocking, habitat enhancements had a contrasted effect on prey communities in mesocosms with coarse woody habitat (CWH) having no significant effect on zooplankton communities but a significant effect on macroinvertebrate community structure while caged-CWH had a significant effect on richness and density of zooplankton and on the richness, density and structure of benthic macroinvertebrate communities. In addition, the presence of caged-CWH was associated with a decrease in the decomposition rate of organic matter. We then found that stocked young-of-the-year largemouth bass induced significant changes in the structure of zooplankton communities, while macroinvertebrate communities and ecosystem functioning were not significantly impacted by stocking. We then found that the impacts of stocking on zooplankton community structure were compensated by habitat enhancement. However, we also found that habitat enhancement with caged-CWH had a significant impact on fish body condition.

While predatory fish stocking has been reported to induce important trophic cascade (Eby et al., 2006; Rudman et al., 2016), we found a significant effect of stocking on zooplankton community structure but limited cascading effect on ecosystem functioning. Specifically, in presence of largemouth bass, we found that the density of some important zooplankton taxa strongly decreased (e.g., Daphnia and Chaoboridae), while the density of some small-bodied cladocerans (e.g., Bosminidae) increased. Such changes in the zooplankton community have already been observed in lakes after largemouth bass introduction (Baca and Drenner, 1995). By consuming large zooplankton taxa, largemouth bass have likely enabled smaller zooplankton to dominate. Surprisingly, we observed an increased phytoplankton with caged-CWH while zooplankton community did not differ from the control without stocking. This might be explained by the fact that food webs in this treatment were already weakened before stocking, with a different community structure and, in particular, a lower richness and density of organisms.

Regarding the benthic habitat, stocking of young-of-the-year bass did not affect, in our experiment, macroinvertebrate community, although they represent potential prey for them (Post, 2003; Brown et al., 2009). The introduction of predatory game fish into fishless lakes has been reported to strongly affect benthic macroinvertebrates (Knapp et al., 2005) and particularly largemouth bass (Jackson, 2002). The lack of effects measured here might be caused by the fact that largemouth bass used in this experiment were reared in a hatchery and fed with pelletized food prior to their introduction in the mesocosms. They may have exhibited different feeding behavior due to rearing conditions (Huntingford, 2004), such as surface feeding with limiting experience in consuming benthic prey. This could also explain why we only observed a consumption effect of largemouth bass on zooplankton without a cascading top-down effect on primary producers.

Contrary to a previous work conducted in a boreal lake (Theis et al., 2022), habitat enhancement did not increase macroinvertebrate productivity in the present study. This result is surprising as the presence of CWH should have provided a prime habitat for benthic macroinvertebrates and a potential refuge from predators (Everett and Ruiz, 1993; Smokorowski et al., 2006; Czarnecka, 2016). A previous study in stream (Bond et al., 2006) showed a very rapid colonization of habitats by macroinvertebrates, within a month, but followed by a drastic drop in abundance and richness after several months. When adding an iron cage with the idea to increase habitat complexity, deleterious effects on the invertebrate community were observed, contrary to previous studies (Schmude et al., 1998; Taniguchi et al., 2003). We found that the presence of cage increased water conductivity (Fig. S10) and this might explain the effect on communities compared to CWH without cage (Cormier et al., 2013). The changes observed in zooplankton diversity and density can also be explained by this change in water quality (Yan et al., 1996). Macroinvertebrate communities found in the two added habitats differed in terms of diversity and structure, indicating that the choice of materials is crucial. We also observed that habitat enhancement did not have positive effects on fish body condition, and we even found significantly lower body mass in fish in the treatment with caged-CWH. This difference can be explained by the fact that the addition of the cage resulted in a significant decrease in the density and richness of zooplankton, thereby reducing the abundance of potential food resources for YOY largemouth bass (Post, 2003; Brown et al., 2009). Specifically, cages might release potential contaminants (e.g., zinc, aluminum) and previous studies have shown that metallic pollution can affect the health and behavior of fish (Javed and Usmani, 2019; Jacquin et al., 2020). It can also have an impact on the structure and functioning of freshwater communities, although its effect can be modulated by other factors such as temperature (Van de Perre et al., 2016). Therefore, further investigations are needed to unsure the absence of ecological effects of the materials used for habitat enhancement (Cooke et al., 2023).

In conclusion, our experience has provided some insights on the capacity of habitat enhancement in potentially limiting the ecological impacts of stocking predatory fish. We have demonstrated that habitat enhancement can partially compensate the predatory effect of fish, but that it can also alter water quality and key ecosystem functions. Ecosystem-based management is promising (Feng et al., 2023; Radinger et al., 2023), but remains complex to implement, particularly in the choice of materials used to ensure a sustainable and beneficial outcome (Cooke et al., 2023). The effects of habitat enhancement occur over the long-term whereas the effects of stocking may occur in the short term, indicating that the impacts of stocking with non-native predators are difficult to reverse, despite investment in habitat enhancement. This study suggests that the outputs of habitat enhancement are complex, and that the use of habitat enhancement alone might not be sufficient to compensate the rapid consequences induced by fish stocking.

Supplementary material

Figure S1. Block design of the experiment with each block (n = 6) containing each of the four treatments.

Figure S2. Additional habitats used in the experiment: (a) coarse woody habitat (CWH) composed of a deadwood bundle (downy oak, Quercus pubescens) and (b) caged-CWH composed of a deadwood bundle (downy oak, Quercus pubescens) installed within an iron cage (Biohut®, ECOCEAN, Montpellier, France).

Figure S3. Initial stocking biomass of largemouth bass Micropterus nigricans in treatments with stocking. For all treatments, five hatchery-reared young-of-the-year (YOY) largemouth bass were introduced in each mesocosm (ntotal = 90 individuals). There was no significant different between treatments in initial stocking biomass (Tukey post-hoc test, p > 0.05).

Figure S4. Density (ind.L−1) of each taxon zooplankton sampled in each treatment (a) before and (b) after fish stocking.

Figure S5. Effect of management practices (a-b) before and (c-d) after fish stocking on zooplankton richness and density (ind.L−1). Different letters indicate significant difference between treatments (Tukey post hoc test, p < 0.05).

Figure S6. Density (ind.mesocosm−1) of sampled macroinvertebrate per taxon and for each treatment (a) before stocking and (b) after fish stocking.

Figure S7. Effect of management practices (a-b) before and (c-d) after fish stocking on macroinvertebrate richness and density (ind.mesocosm−1). Different letters indicate significant difference between treatments (Tukey post hoc test, p < 0.05).

Figure S8. Density (ind.hab−1) of each taxon of macroinvertebrates for each treatment with a habitat enhancement device (CWH only and caged-CWH) after fish stocking.

Figure S9. Macroinvertebrate (a) richness and (b) density (ind.hab−1) within the CWH and the caged-CWH. Different letters indicate significant difference between treatments (ANOVA, p < 0.05).

Figure S10. Effects of treatments (a) before and (b) after stocking on conductivity (μS.cm−1). Different letters indicate significant difference between treatments (Tukey post hoc test, p < 0.05).

Tab. S1. Details on the models, transformation and distributions used for each response variable before and after stocking.

Access here

Acknowledgements

We are grateful to S. Boutes for access to the experiment site, to the Gravière Team for their help during the experiment and to ECOCEAN for providing Biohut®. This research was funded in France by the OFB (Office Français de la Biodiversité) and the Région Occitanie as part of the SPECTRA project. This study is part of the long-term Studies in Ecology and Evolution (SEE-Life) program of the CNRS.

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Cite this article as: Gimenez M, Cucherousset J. 2024. Can habitat enhancement limit the ecological impacts of predatory fish stocking? An experimental approach. Knowl. Manag. Aquat. Ecosyst., 425, 8.

All Figures

thumbnail Fig. 1

Effects of management practices before stocking on community structure quantified using (a-b) a non-metric multidimensional scaling (nMDS) analysis and analyzed along the (c-d) first and (e-f) second axes. Zooplankton communities are represented on the left side of the panel and macroinvertebrate communities on the right side. Different letters indicate significant difference between treatments (Tukey post hoc test, p < 0.05).
In the text
thumbnail Fig. 2

Ecological effects of treatments before stocking on (a) pelagic primary production, (b) benthic primary production and (c) decomposition rate. Different letters indicate significant difference between treatments (Tukey post-hoc test, p < 0.05).
In the text
thumbnail Fig. 3

Treatment effects on (a) scaled-mass index and (b) survival rate of juvenile largemouth bass Micropterus nigricans, at the end of the experiment. Different letters indicate significant difference between treatments (Tukey post-hoc test, p < 0.05).
In the text
thumbnail Fig. 4

Effects of management practices after stocking on community structure quantified using (a-b) a non-metric multidimensional scaling (nMDS) analysis and analyzed along the (c-d) first and (e-f) second axes. Zooplankton communities are represented on the left side of the panel and macroinvertebrate communities on the right side. Different letters indicate significant difference between treatments (Tukey post hoc test, p < 0.05).
In the text
thumbnail Fig. 5

Macroinvertebrate community structure within the added habitats after stocking quantified using (a) a non-metric multidimensional scaling (nMDS) analysis and analyzed with the (b) first and (c) second axes. Different letters indicate significant difference between treatments (ANOVA, p < 0.05).
In the text
thumbnail Fig. 6

Ecological effects of management practices after stocking on (a) pelagic primary production, (b) benthic primary production and (c) decomposition rate. Different letters indicate significant difference between treatments (Tukey post-hoc test, p < 0.05).
In the text

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