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All issues No 425 (2024) Knowl. Manag. Aquat. Ecosyst., 425 (2024) 13 Full HTML
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Knowl. Manag. Aquat. Ecosyst.
Number 425, 2024
Anthropogenic impact on freshwater habitats, communities and ecosystem functioning
Article Number 13
Number of page(s) 9
DOI https://doi.org/10.1051/kmae/2024009
Published online 12 July 2024

© L. Wang et al., 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

Aquatic species have been extensively introduced by humans, intentionally or unintentionally, through trade (Novák et al., 2020), aquaculture (Haubrock et al., 2021), canals (Hulme, 2009) and some area considered invasive to the region. Biological invasions are among the most serious threats to ecosystems and such invasions can lead to extinction of native species, loss of biodiversity and alteration of ecosystem functions (Vitousek et al., 1997; Ewel et al., 1999; Mack et al., 2000; Sala et al., 2000). Negative impacts of invasive species on aquatic systems have been extensively reported (Carlsson and Lacoursiere, 2005; Matsuzaki et al., 2009; Jackson et al., 2017; Britton, 2023).

The golden apple snail Pomacea canaliculata (Lamarck) is an aquatic species native to South America and it is listed as one of the top 100 worst invasive alien species in the world by the International Union for the Conservation of Nature (Lowe et al., 2000). Because of its high reproductive rate, rapid growth, high adaptability, and few natural enemies, P. canaliculata has spread rapidly to many natural aquatic habitats (Naylor, 1996; Lach and Cowie, 1999). Nowadays, it is widely distributed and continues to expand its range (Baker, 1998; Carlsson and Lacoursiere, 2005; Lv et al., 2011; Yin et al., 2022).

Both the juvenile and adult individuals can feed on macrophytes (Lach et al., 2000; Cowie, 2002; Boland et al., 2008; Qiu and Kwong, 2009; Kwong et al., 2010). The species is more selective for submerged plants than emergent plants (Estebenet and Martín, 2002; Carlsson and Brönmark, 2006; Wang and Pei, 2012). Submerged macrophytes, with higher nutritional value and lower defenses, are ideal food for P. canaliculata (Qiu and Kwong, 2009) in comparison with other macrophytes, reflecting that nutrient content, physical structure (tough leaves, etc.) and dry matter content of macrophytes influence their palatability (Elger and Lemoine, 2005; Boland et al., 2008; Qiu and Kwong, 2009; Wong et al., 2010; Qiu et al., 2011). Snails can therefore affect the macrophyte species assemblage and it is known to greatly reduce the overall biomass of macrophytes in some systems (Carlsson and Lacoursiere, 2005). A survey of natural wetlands in Thailand showed an almost complete disappearance of macrophytes in areas with high densities of this snail (Carlsson et al., 2004).

Submerged nmacrophytes are essential for maintaining a high-quality ecosystem state in shallow lakes (Jeppesen et al., 1998; Van Donk, 1998). They increase habitat complexity and provide shelter for zooplankton and macroinvertebrates (Timms and Moss, 1984; Brönmark and Vermaat, 1998), absorb nitrogen and phosphorus from the water column (Ozimek et al., 1990; Kufel and Ozimek, 1994), often release allelochemicals that inhibit phytoplankton growth (Jasser, 1995), and enhance sedimentation by reducing resuspension (Petticrew and Kalff, 1992; Barko and James, 1998). In addition, submerged macrophytes are important primary producers in shallow lakes (Jin et al., 2020), providing food for various aquatic animals such as macroinvertebrates and herbivorous fish. Submerged macrophyte restoration has been a key method for remediating eutrophic waters (Liu et al., 2018; Xue et al., 2020).

The existing literature documenting the effects of P. canaliculata on aquatic ecosystems have focused mostly on submerged macrophytes (Fang et al., 2010; Wang and Pei, 2012; Liu et al., 2021; Gao et al., 2021; Yang et al., 2021) and phytoplankton (Carlsson et al., 2004). Few studies have simultaneously investigated the effect of invasive species on the location of primary production in shallow lakes (Martín et al., 2019). Benthic and littoral macrophytes and phytoplankton share light and nutrient requirements but grow in different but coupled habitats. A shift of primary production from benthic macrophytes to open water phytoplankton can lead to serious ecological problems.

We hypothesised that the consumption of benthic primary producers by P. canaliculata will inhibit the growth of submerged macrophytes, increase nutrient concentrations in the water column, and thus enhance phytoplankton growth. Therefore, P. canaliculata will shift primary production from benthic to pelagic habitat resulting in deteriorating water quality. To test this hypothesis, we conducted a short-term experiment with and without P. canaliculata in mesocosms dominated by Vallisneria natans (Lour.) Hara. This submerged macrophyte was used because it is perennial, rooted and widely distributed in freshwater lakes, ponds and rivers in China and around the world (Xie et al., 2005; Li et al., 2018). V. natans exhibits better survival and growth potential than other submerged plants in eutrophic lakes (Qiu et al., 2001) and is currently the most popular aquatic plant used for ecological restoration in China (Wang et al., 2011).

2 Materials and methods

2.1 Experimental mesocosm set-up

Eight circular tanks (upper diameter 40 cm, lower bottom diameter 33 cm and height 41 cm) were used as mesocosms holding sediment, water and Vallisneria natans. The sediment (TN = 3.40 mg · g−1; TP = 1.52 mg · g−1) were collected from Ming Lake (23°13'70”−23°13'75” N, 113°35'62”−113°35'74” E), a shallow eutrophic lake in Guangzhou, China, air dried and then sieved through a 0.5 mm mesh sieve to remove coarse particulates and debris as well as large-bodied macroinvertebrates. The water was from Ming Lake and filtered (TN = 1.58 mg · L−1; TP = 0.09 mg · L−1) on a plankton net (pore size, 64 µm) to remove large zooplankton to avoid unnaturally high grazing pressure on the phytoplankton as fish were absent due to the small scale of the experiment. We added a 5 cm thick layer of sediment to each mesocosm before filling them with water. Thereafter we planted 25 V. natans (wet weight 250 ± 2 g and leaf length 35 ± 2 cm) and kept the mesocosms for two weeks under outdoor natural sunlight conditions before the start of the experiment. P. canaliculata were purchased from a market in Guangzhou, kept for 24 h in a tank with lake water without food before introducing them to the mesocosoms. One individual of P. canaliculata (average shell height 4 ± 0.2 cm and fresh weight 12 ± 0.5 g) was added to each of four mesocosms as snail treatment. The other four mesocosms served as control. During the experiment, lake water was added when needed to maintain water level in the mesocosm. The mesocosms were not covered and the experiment was carried out at natural sunlight from May 17 to July 1, 2022 and the air temperature was 17−35 °C during the experiment.

2.2 Sampling and analysis

Every nine days, water samples (1 L) were collected from 20 to 30 cm below the water surface of each mesocosm using clean polyethylene bottles. The water samples were used for measurements of total nitrogen (TN), nitrate nitrogen (NO3−-N), ammonium nitrogen (NH4+–N), total phosphorus (TP), soluble reactive phosphorus (SRP), nitrogen to phosphorus ratio (TN:TP), total suspended solids (TSS) and organic suspended solids (OSS).

Phytoplankton biomass was measured as size-fractionated chlorophyll for microphytoplankton (>20 μm), nanophytoplankton (2–20 μm), and picophytoplankton (0.2–2 μm). For size fractionation, 200 mL of water was filtered on a 20 μm filter (nylon mesh) for microphytoplankton, then a 2 μm filter (nylon mesh) for nanophytoplankton, and finally a 0.2 μm filter (nylon mesh) for picophytoplankton biomass (Rong et al., 2021). Chl a of phytoplankton was extracted with 90% acetone for 24 h and Chl a was measured spectrophotometrically (MEP of PRC, 2017). The total phytoplankton biomass was calculated as the sum of the biomass of the three components. Nutrients were measured according to the American Public Health Association (APHA, 1998). A 400 mL water sample was filtered on a Whatman GF/C fiber membrane, dried at 108 °C for 2 h, weighed to obtain TSS, and then ashed at 550 °C for 2 h to obtain data for organic suspended solids (OSS) (Huang et al., 1999).

At the end of the experiment, V. natans was collected from each mesocosm, washed and then wet weight, leaf length and tiller numbers (branching of plants occurring below or near the ground) were determined. The plants were then dried at 80 °C to constant weight for dry weight determination. The wet weight and shell height of each P. canaliculata were also recorded at the end of the experiment.

2.3 Statistical analyses

Repeated measures analysis of variance (RM-ANOVAs) was used to analyze the differences in nutrient concentrations, biomasses of phytoplankton in different sizes, TSS and OSS between the snail treatment and the control, with time as the repeated factor. Prior to analysis, Mauchly's test of sphericity and test of homogeneity of variances were performed to meet the assumption. Independent-samples t-test was performed to compare differences of wet weight, tiller number, and leaf length of V. natans between the control and the snail treatment at the end of the experiment. Prior to analysis, the assumption of equality of variances was tested with Levene's test. Statistical analyses were carried out using IBM SPSS Statistics 25.0 software. All data were presented as mean values ± 1SD.

3 Results

3.1 Nitrogen

Concentrations of TN and NO3−-N were higher in the snail treatment than in the control (RM-ANOVAs, treatment effect, p = 0.001 and 0.016, respectively, Fig. 1), while NH4+-N was not significantly different (p = 0.114). TN, NO3−-N and NH4+-N varied over time (RM-ANOVAs, time effect, p = 0.002, 0.004 and 0.007, respectively). The interaction effects of time and treatment for TN and NO3−-N were significant (RM-ANOVAs, interaction effect, p = 0.012 and 0.012, respectively), but not for NH4+-N (p = 0.175).

thumbnail Fig. 1

Nitrogen concentration (TN, NO3−-N, NH4+-N, mean ± 1SD) in the control and the snail treatment over time.

3.2 Phosphorus and TN:TP

The concentrations of SRP were lower in the snail treatment than in the control (RM-ANOVAs, treatment effect, p = 0.002, Fig. 2), while TP and TN:TP ratio were not significantly different (RM-ANOVAs, treatment effect, p = 0.216 and 0.182). TP, SRP and N: P ratio varied with time (RM-ANOVAs, time effect, each p < 0.001). The interaction effects between treatment and time for TP, SRP and TN:TP ratio were significant (RM-ANOVAs, interaction effect, p = 0.012, < 0.001 and 0.030, respectively).

thumbnail Fig. 2

Phosphorus concentrations (TP, SRP and TN:TP, mean ± 1 SD) in the control and the snail treatment over time.

3.3 Total suspended solids (TSS) and organic suspended solids (OSS)

The concentrations of TSS and OSS were higher in the snail treatment than in the control (RM-ANOVAs, treatment effect, p = 0.012 and 0.001, respectively, Fig. 3) and varied with time (RM-ANOVAs, time effect, p = 0.001 and 0.001, respectively). The interaction effects for TSS and OSS were significant (RM-ANOVAs, interaction effect, p = 0.001 and 0.001, respectively).

thumbnail Fig. 3

TSS and OSS concentrations in the water (mean ± SD) in the control and the snail treatment over time.

3.4 Biomass of phytoplankton and size-fractionated phytoplankton

The biomass of phytoplankton (Chl a) was higher in the snail treatment than in the control (RM-ANOVAs, treatment effect, p = 0.034, Fig. 4). The Chl a of phytoplankton varied with time (RM-ANOVAs, time effect, p = 0.048). The interaction effects for Chl a of phytoplankton were significant (RM-ANOVAs, interaction effect, p = 0.017).

The microphytoplankton biomass (Chl a) was higher in the snail treatment than in the control (RM-ANOVAs, treatment effect, p = 0.002, Fig. 5), while the biomasses (Chl a) of nanophytoplankton and picophytoplankton did not differ among treatments (RM-ANOVAs, treatment effect, p = 0.083 and 0.278, respectively). The microphytoplankton biomass (Chl a) varied with time (RM-ANOVAs, time effect, p < 0.001), while the biomasses (Chl a) of nanophytoplankton and picophytoplankton did not (p = 0.203 and 0.204, respectively). The interaction effects for microphytoplankton biomass (Chl a) were significant (RM-ANOVAs, interaction effect, p <0.001), but not for the biomasses (Chl a) of nanophytoplankton and picophytoplankton (p = 0.383 and 0.095, respectively).

thumbnail Fig. 4

Chl a of phytoplankton and size-fractionated phytoplankton (microphytoplankton, nanophytoplankton and picophytoplankton) in the control and the snail treatment over time.
thumbnail Fig. 5

Wet weight, leaf length, tiller number (mean ± 1 SD) of Vallisneria natans in the control and the snail treatment.

3.5 Wet weight, leaf length and tiller number of Vallisneria natans

Wet weight, leaf length and tiller number of V. natans were lower in the snail treatment than in the control (t-test, p< 0.001, < 0.001, = 0.042, respectively, Fig. 5).

4 Discussion

We found that the presence of P. canaliculata increased TN, NO3−-N, TSS and OSS concentrations, while we did not find any change in TP and TN:TP, but a decrease in SRP. We also found that P. canaliculata increased the total phytoplankton biomass and microphytoplankton biomass but did not alter the biomass of nanophytoplankton and picophytoplankton. The biomass, leaf length and tiller number of V. natans also decreased markedly in the snail treatment.

The increase in nitrogen concentrations of TN and NO3−-N in the snail treatment likely derives from the excretion by P. canaliculata (Pan et al., 2014). In addition, the consumption by the animal may also release nitrogen from the submerged macrophyte to the overlying water. Furthermore, the bioturbation of the animal on sediments when moving may enhance the release of nitrogen to overlying water. However, we cannot discount the possibility that some atmospheric nitrogen was added to the water by phytoplankton fixation as the P. canaliculata did change the phytoplankton community. Unfortunately, we did not monitor changes in the phytoplankton species composition.

We found a major increase in phosphorus of TP and SRP in both the treatment and the control during the course of the experiment but no difference in TP between them. This pattern can only be explained by a major release of phosphorus from the sediment perhaps associated with the increased temperatures in the mesocosms with the experiment running in summer from May to July, as the release rate increases with increasing temperature (Nicholls, 1999; Søndergaard et al., 2003). However, the concentrations of SRP decreased in the snail tanks compared to the controls, likely reflecting a higher uptake in phytoplankton due to their higher biomass here than in the control. Nitrogen and phosphorus are often limiting nutrients for phytoplankton growth (Elmgren and Larsson 2001; Bledsoe et al., 2004). Pomacea canaliculata can increase both the nitrogen and phosphorus concentrations (Carlsson et al., 2004). We did find that the snail increased total nitrogen in this study, but we did not find any change in total phosphorus and TN: TP.

We also found a shift in phytoplankton size structure as only the microphytoplankton increased in the snail treatment while the smaller size fractions were not affected. Although debris particles, phytoplankton, invertebrates and macrophytes can be food sources for P. canaliculata (Kwong et al., 2009; Wong et al., 2009; Kwong et al., 2010), its primary food is macrophytes (Carlsson et al., 2004; Carlsson and Lacoursiere, 2005). The changes in phytoplankton size structure is therefore not likely to be directly related to feeding, but instead due to effects of difference in the nutrient concentrations between the snail treatment and the control, or to changes in the light condition. However, the changes in phytoplankton size structure can lead to profound change of ecosystem, including food web, efficiency of material transformation, and so on.

With the enhanced growth of phytoplankton in the snail treatment, the TSS and OSS increased thereby increasing the light attenuation and decreasing the light intensity at the sediment surface which may limit the growth of V. natans. Thus, P. canaliculata can negatively affect the growth of V. natans directly by consuming the plant, and indirectly by increasing the light attenuation. We did find that the biomass, leaf length and tiller number of V. natans were lower in the P. canaliculata treatment than in the control, indicating that the snail hampered the growth of the submerged macrophytes. Similarly, Calvo et al. (2019) showed that consumption by P. canaliculata can slow the growth of submerged macrophyte (Hydrilla verticillata (L. f.) Royle). However, whether the reduction of submerged macrophyte growth is caused by direct consumption of the plant, by shading, or both cannot be determined from our study.

Our results need to be interpreted with caution. The results may be used to assess specific mechanisms of primary production shifting from benthic to pelagic habitats by the snail P. canaliculata but may not accurately represent the variations occurring in natural lakes due to scale differences. However, the results coincide with other studies showing that the snail can reduce the growth of submerged plants, increase the turbidity and deteriorate the water quality not only in mesocosms (Liu et al., 2021; Gao et al., 2021; Yang et al., 2021), but also in natural lakes (Martín et al., 2019).

In summary, the P. canaliculata increased TN, NO3−-N, TSS and OSS concentrations, did not change TP and TN: TP, but decreased SRP, enhanced the growth of phytoplankton and microphytoplankton, altered the phytoplankton communities, and decreased the light intensity and the growth of submerged macrophyte. Thus, P. canaliculata shifted primary production from benthic macrophytes to pelagic phytoplankton, altered the phytoplankton community composition and deteriorated water quality. Removal or control of P. canaliculata would be useful to protect submerged macrophyte and maintain clear water in lakes and ponds.

Acknowledgements

This work was supported by the National Key Research and Development Program of China (No. 2022YFE0122100). EJ was supported by the TÜBITAK program BIDEB2232 (project 118C250). We acknowledge comments by Professor Mariana Meerhoff to an earlier draft of this manuscript.

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Cite this article as: Wang L, Ren L, Gao B, Jeppesen E, Rudstam LG, Karpowicz M, Feniova I, Liu Z, Tang Y, Zhang X. 2024. The golden apple snail Pomacea canaliculata shifts primary production from benthic to pelagic habitats in simulated shallow lake systems. Knowl. Manag. Aquat. Ecosyst., 425. 13

All Figures

thumbnail Fig. 1

Nitrogen concentration (TN, NO3−-N, NH4+-N, mean ± 1SD) in the control and the snail treatment over time.
In the text
thumbnail Fig. 2

Phosphorus concentrations (TP, SRP and TN:TP, mean ± 1 SD) in the control and the snail treatment over time.
In the text
thumbnail Fig. 3

TSS and OSS concentrations in the water (mean ± SD) in the control and the snail treatment over time.
In the text
thumbnail Fig. 4

Chl a of phytoplankton and size-fractionated phytoplankton (microphytoplankton, nanophytoplankton and picophytoplankton) in the control and the snail treatment over time.
In the text
thumbnail Fig. 5

Wet weight, leaf length, tiller number (mean ± 1 SD) of Vallisneria natans in the control and the snail treatment.
In the text

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