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Article

Impact of Grass Carp and Crucian Carp on Submerged Macrophyte and Phosphorus Cycling in Shallow Lake Mesocosms

by
Xin Chen
1,2,
Weiju Wu
3,
Hongyi Ao
1,
Shenghua Hu
4,
Huaqiang Chen
4,
Xiaofei Chen
3 and
Chenxi Wu
1,*
1
Institute of Hydrobiology, Chinese Academy of Sciences, Wuhan 430072, China
2
University of Chinese Academy of Sciences, Beijing 100039, China
3
Hubei Provincial Academy of Eco-Environmental Sciences, Wuhan 430072, China
4
Wuhan Bridge Engineering Co., Ltd., Wuhan 430061, China
*
Author to whom correspondence should be addressed.
Water 2025, 17(3), 326; https://doi.org/10.3390/w17030326
Submission received: 24 December 2024 / Revised: 21 January 2025 / Accepted: 21 January 2025 / Published: 24 January 2025
(This article belongs to the Section Biodiversity and Functionality of Aquatic Ecosystems)
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Abstract

:
Submerged macrophytes are essential for the restoration of shallow lakes for maintaining clear-water conditions. The presence of fish can affect the nutrient cycles and the growth of submerged macrophytes in lakes. In this study, a 28-day mesocosm experiment was carried out with an herbivorous fish Ctenopharyngodon idella (CID) and an omni-benthivorous fish Carassius auratus (CAU) to investigate their effects on the growth of a submerged macrophyte Hydrilla verticillata and phosphorus (P) cycle in shallow lakes. The results showed that CID slowed down the growth of H. verticillata while CAU showed no significant effect. In overlying water, CID only increased the ammonium nitrogen (NH3-N) concentration in the later stage due to excretion, while CAU elevated particulate phosphorus (PP) levels during the experiment through disturbance. Meanwhile, the radial oxygen loss and photosynthesis of H. verticillata in CAU might promote the formation of NaOH-P and HCl-P in the sediment, respectively. Changes in the water and sediment properties caused by CID and CAU can contribute to the increase in the eutrophication risk index (ERI). Our findings suggest that CID has the potential to be an indirect biological manipulation tool, while CAU should be controlled to minimize its negative impacts on the P cycle in lakes.

Graphical Abstract

1. Introduction

Eutrophication is one of the most widespread aquatic environmental issues globally under the combined impacts of anthropogenic and natural factors, which may result in the subsequent dominance of phytoplankton and the disappearance of submerged macrophytes [1,2]. The restoration of submerged macrophytes is one of the most effective methods for ecological remediation of many eutrophic shallow lakes [3]. Its success depends on conditions such as light, the properties of the water column and sediment, and bioturbation by fish [4,5].
Herbivores (herein, we refer to herbivores that feed on submerged macrophytes) can negatively impact the recovery of submerged macrophytes via a top-down effect [6]. They assimilate the nutrients from submerged macrophytes or convert them into feces. Besides their consumptive role, some positive effects of herbivores on submerged macrophytes have been observed, such as promoting over-compensatory growth and halting the regime shift in shallow lakes from clear to turbid after the restoration of submerged macrophytes [7,8]. However, the effect of herbivorous fish activities on the overall nutrient dynamics between overlying water and sediment during the restoration process has received limited attention.
The effects of herbivores on submerged macrophytes are related to the feeding habits of herbivores and the growth strategy of submerged macrophytes. Dominant omni-benthivorous fish in subtropical lakes, such as crucian carp (Carassius auratus, CAU), regard submerged macrophytes as part of their food source, while their feeding rate is lower than that of typical herbivorous, such as Ctenopharyngodon idella (CID) and Parabramis pekinensis [9]. In addition, fish behaviors such as disturbance and defecation can change nutrient concentration in the water column, which in turn can indirectly affect the growth of plants that follow a fast growth strategy, such as Hydrilla verticillata and Myriophyllum spicatum, in the short term [10]. Nevertheless, studies on the effects of fish with different feeding habits on the growth of submerged macrophytes are limited.
Phosphorus (P) is the primary driver of eutrophication in many shallow lakes [11]. Consequently, this study aims to explore the dynamics of P processes in overlying water and sediments using mesocosm experiments with the following objectives: (1) elucidate the effects of fish with different feeding habits on the growth of submerged macrophytes; (2) reveal the process and mechanism of P cycling during the experiment; and (3) provide a theoretical foundation for the regulation of fish after submerged macrophytes restoration in shallow lakes.

2. Materials and Methods

2.1. Experimental Setup

The experiment was conducted from 14 May to 11 June 2024 at the field research station located on the shore of a pond next to Liangzi Lake in Hubei Province (30°32′56″ N, 114°21′13″ E, Figure S1). A total of 12 high-density polyethylene containers (inner diameter at the top: 120 cm; inner diameter at the bottom: 90 cm; height: 91 cm; volume: 700 L) were used for the experiment. The situ sediments were collected from the pond and processed to remove large debris and then homogenized and layered at the bottom of each container to a height of 10 cm. The in situ overlying water from the pond was filtered through a phytoplankton net with 64 µm mesh and was then injected into each container, amounting to 600 L. The total nitrogen (TN), total phosphorus (TP), and chlorophyll a (Chl. a) contents (mean ± standard error) of the in situ water were 0.72 ± 0.09 mg/L, 0.031 ± 0.004 mg/L, and 4.81 ± 1.65 µg/L, respectively. The TP and organic matter (OM) contents of the sediment samples were 0.733 ± 0.21 mg/g and 68.61 ± 3.64 mg/g, respectively.

2.2. Plant and Fish Species

Hydrilla verticillata (H. verticillata), a typical submerged macrophyte widely distributed in the middle and lower reaches of the Yangtze River Basin, was used for the experiment [12]. H. verticillata was purchased from a nursery near the research station and was originally planted in the lake. Top 20 cm apical branches of H. verticillata were clipped, then rinsed with water to exclude the attached algae, and pre-cultivated with in situ water at the experiment station for spare [13].
Typical herbivorous fish CID and omni-benthivorous fish (CAU) were used as the experimental fish and were purchased from Yongchuan Fishery, Chongqing, China. Both species have been demonstrated to have significant influences on lake ecosystems in the middle-lower Yangtze River Basin [14,15]. Before the experiment, all fish were temporarily reared in water from the pond to adjust to the experimental conditions.

2.3. Experimental Design

The control group (no fish, NF) was planted with only H. verticillata. The treatment groups were planted with H. verticillata and stocked with either CID or CAU. Each treatment group included four replicates. On 30 April 2024, H. verticillata was planted into the 12 containers, with the initial total biomass in each container being 53.08 ± 0.30 g/m2. After 14 days of pre-cultivation, 1 grass carp (wet weight: 10.89 ± 1.59 g, total length: 11.38 ± 1.18 cm) or 1 crucian carp (wet weight: 9.76 ± 1.87 g, total length: 9.38 ± 0.25 cm) were introduced to each container, and the initial density of fish was 16.23 ± 1.72 g/m2 (1.57 ind./m2).

2.4. Sampling and Analysis

2.4.1. Overlying Water

During the experiment, 2.0 L water samples were collected 20 cm below the water surface every 7 days for the analysis of water quality. Water temperature (WT), pH, dissolved oxygen (DO), and oxidation/reduction potential (ORP) of the water were measured in situ using a multiparameter analyzer (HQ40d, Hach, Loveland, CO, USA). TP, TN, nitrite nitrogen (NO3-N), and ammonium nitrogen (NH3-N) were determined following standard methods (GB 11893-89 [16], HJ 636-2012 [17], HJ/T 346-2007 [18], and HJ 535-2009 [19]). Total dissolved phosphorus (TDP) and soluble reactive phosphorus (SRP) were determined by referring to TP after water sample filtration and filtration without digestion, respectively. Particulate phosphorus (PP) was obtained by TP minus TDP. Total dissolved nitrogen (TDN) was measured the same as TN after water sample filtration. Chl. a was measured spectrophotometrically following extraction with 90% acetone. Total suspended solids (TSSs) were determined gravimetrically as matter retained on the GF/C filter after drying at 105 °C for 24 h. The filters were then baked in a muffle furnace at 550 °C for 2 h for the analysis of inorganic suspended solids (ISSs).

2.4.2. Sediment

The surface sediments were collected at the start and the end of the experiment for the determination of TP, P fractions, pH, and OM. The content of TP in sediment was determined by the alkali fusion method. P fractions in the sediment were extracted by the modified SMT protocol [20], with defined scheme comprising Ex-P (loosely exchangeable P), NaOH-P (Fe/Al-P, P bound to Fe, Al, and Mn oxides and hydroxides), HCl-P (Ca-P, P associated with apatite), inorganic P (IP), and organic P (OP). The procedure is provided in Supplementary Materials (Figure S2). Briefly, 0.3 g of freeze-dried sediment was vibrated and extracted with different chemical extractants step-by-step at a constant solid/solution ratio (1:100). The pH of the sediment was measured using a pH meter after mixing the sediment sample with deionized water at a ratio of 1:2.5 [21]. The OM content of the sediment was determined using the weight loss-on-ignition method [22]. The equilibrium P concentration (EPC0) was obtained by the Freundlich model fitting equation [23], and the detailed method is presented in Supplementary Materials (Text S1). The eutrophication risk index (ERI, %) was used to assess the risk of eutrophication, which is determined by the ratio of P saturation level (DPS, %) and the P sorption index (PSI, [mg P/(100 g)]/[μmol/L]), referring to the method of [24].

2.4.3. Plankton Community

At the beginning and the end of the experiment, 1 L depth-integrated water samples were taken from each container. These samples were added to 10 mL of Lugol’s iodine solution for later analysis of phytoplankton after a 48-h settling period. Samples were concentrated by siphon method, counted, and then taxonomically identified to the species level by referring to “The Freshwater Algae of China: systematics, Taxonomy and Ecology” [25].

2.5. Data Analysis

Figures were drawn by Origin Pro 2024 software and statistical analysis was performed using the SPSS 25 software. Error bars represent the standard deviations calculated from four replicate measurements in each figure. The significant differences between time and among treatments were analyzed using a t-test and one-way analysis of variance (ANOVA); differences in treatments during the experiment were analyzed using two-way ANOVA. Spearman correlation analysis was used to reveal the relationships among environmental parameters at the p < 0.05, p < 0.01, and p < 0.001 levels of significance. A partial least squares path model (PLS-PM) was employed to evaluate the interrelationships among water properties, phytoplankton, nutrients in water, sediment properties, P in sediment, and sediment P release. The analysis was performed using the “plspm” package in R studio 4.0.4 software [26]. The relationship between latent variables was established based on hypothesized causal pathways informed by previous studies [27].

3. Results

3.1. Changes in the Biomass of Fish and H. verticillata

All fish survived at the end of the 28-day experiment. The biomass of CID increased significantly by 163.87%, while that of CAU decreased by 2.56% (Figure 1A). The biomass of H. verticillata in CID, CAU, and NF (Figure 1B) had significantly increased by 9.78, 35.53, and 34.70 times compared to the initial stage, respectively. The introduction of CID resulted in a significantly lower biomass of H. verticillata than that of CAU and NF groups (p < 0.05).

3.2. Changes in the Overlying Water Properties

The average WT ranged from 22.48 °C to 31.9 °C and was similar in each container during the experiment. Both pH and DO (Figure 2A,B) in the NF group were significantly higher than those in the CID and CAU on day 21, while those in CAU were also significantly higher than CID on day 28. At the end of the experiment, DO was markedly lower in both CID and CAU compared to the initial values.
TSS and ISS (Figure 3A,B) in CAU showed the greatest increase and were higher than CID and NF from day 7 onwards. ISS was the dominant form of TSS in both CID and CAU, accounting for 56.22–83.86%. In both CID and CAU, TP and PP (Figure 3C,D) increased gradually and significantly at the end. They were significantly higher in CAU than other groups from day 7, with a peak on day 14 (TP: 0.194 mg/L, PP: 0.178 mg/L), and then declined, but were still significantly higher at the end of the experiment than the initial values. PP remained the main form of P in the overlying water, constituting 57–99.6% of TP. Throughout the experiment, SRP (Figure 3E) did not show significant differences between groups and among time (two-way ANOVA, p > 0.05).
During the experiment, TN, TDN, and NO3-N (Figure 3F–H) did not present significant differences among the treatments but all showed an overall increasing trend and were significantly higher than the initial values at the end. NH3-N in CID was significantly higher than that in CAU and NF on day 28 with a peak value of 0.40 mg/L.

3.3. Changes in Chl. a and Phytoplankton Community

Chl. a (Figure 4A) was significantly higher in CAU on day 7. By day 28, its level in CID had increased, eliminating the significant difference between the two groups. A total of 75 phytoplankton species were recorded and no significances were observed in either the total abundance or biomass of phytoplankton among the three treatments at the beginning and end of the experiment. Initially, the total phytoplankton abundance and biomass averaged 1.4 × 106 cells/L and 0.64 mg/L, respectively. The phytoplankton community was comprised primarily of cyanophytes (accounting for 70.67% and 9.31% of total abundance and biomass, respectively), cryptophytes (14.70% and 60.44%), and chlorophytes (9.23% and 10.01%) (Figure 4B,C). Ultimately, CID showed the greatest increase in the total abundance (2.4 × 107 cells/L) and biomass (2.37 mg/L) of phytoplankton. In terms of relative abundance, chlorophyte was the largest proportion (76.94%) in CID, while cyanophyte was significantly dominant in CAU and NF, with a proportion of 60.97% and 57.68%, respectively. As for the relative biomass, chlorophyte prevailed in all three treatments (accounting for 72.18% in CID, 48.43% in CAU, and 74.61% in NF). The biomass of cryptophyte in CID was significantly higher than that of NF, while other phytoplankton did not differ significantly among the treatments.

3.4. Changes in Sediment Properties

There were no significant differences in sediment pH (Figure 5A) among treatments and across the time scale, with an average value ranging from 5.85 to 6.29. The OM (Figure 5B) in CID significantly increased, with an average rise of 7.12 mg/g. However, no significant difference was present in the OM among the treatments at the end. The order of the average content of P fractions (Figure 5C–H) in sediment was NaOH-P > OP > HCl-P > Ex-P. IP was the main form of TP in the sediments, of which NaOH-P was the major component. Only CAU showed significant changes in NaOH-P and HCl-P on day 28, with an increase of 0.128 mg/g and 0.018 mg/g, respectively.
The EPC0 and ERI (Figure 6A,B) showed the same significant trend: CAU > CID > NF. The ERI values in both the CID and CAU groups exceeded the threshold of 25%, which is a high-release risk, while that in NF remained below 10%, which is a low-release risk.

3.5. Environmental Drivers of the P Cycle

In order to investigate the effect of physicochemical properties and P component of the overlying water and sediment on sediment P release, correlation (Figure S3) and PLS-PM (Figure 7, Table S1) analyses of these parameters were conducted. After fish activity, TP in overlying water was significantly and positively correlated with Chl. a (R2 = 0.68, p < 0.05), TSS (R2 = 0.68, p < 0.05), and ISS (R2 = 0.70, p < 0.05). Among the sediment parameters, the TP in sediment and ERI was significantly and positively correlated with OM (R2 = 0.67, p < 0.05) and total phytoplankton biomass (R2 = 0.61, p < 0.05), respectively.
PLS-PM revealed the pathways of fish to influence P cycling. The effects of water properties and phytoplankton on nutrients in water were significant, with direct effects of 0.78 (p = 0.003) and −0.61 (p = 0.012), respectively. In addition, there were three main pathways of influence on sediment P release. First, P fractions in sediment were regulated by water properties (direct effect 0.97, p = 0.040), which resulted in an indirect effect (total effect 0.70) on sediment P release. Second, changes in sediment properties (mainly pH) and P fractions (mainly Ex-P, NaOH-P, HCl-P, and OP) could directly affect sediment P release, with the direct effect being 0.67 (p = 0.027) and 0.85 (p = 0.023), respectively.

4. Discussion

4.1. Impact of Fish on H. verticillata Growth

CID consumed a significant quantity of H. verticillata, part of which was used to accumulate nutrients for their growth and part of which was transferred to feces. Also, the photosynthesis of H. verticillata slowed down accordingly with the decrease in biomass, such that pH and DO were lower in CID at the end of the experiment.
The final biomass of H. verticillata in CAU was the highest, which was mainly attributed to the release of nutrients by CAU during disturbance. The fast growth strategy of H. verticillata thus led to rapid nutrient uptake and growth. In contrast, for slow-growing Vallisneria natans, small CAU could inhibit its growth [28]. Since CAU is omni-benthivorous, its impact on TSS in water is greater. Suspended particulates attached to the surface of plant leaves can interfere with CO2 exchange [29]. At the same time, the gradual growth of H. verticillata towards the water surface forms a dense canopy, which limits photosynthesis in the lower H. verticillata layer. The combination of these two reasons resulted in a significantly lower DO in CAU than in NF on day 28, although it was still higher than that in CID with a lower biomass of H. verticillata.

4.2. Impact of Fish on Water Quality

The influences of CID and CAU on nitrogen (N) and P in the overlying water are different. ISS, as a reflection of the strength of fish disturbance of sediments, shows no significant difference between CID and NF, indicating that CID had a weak ability to disturb sediments. Though CID did not significantly affect P fractions in water, it contributed to significantly higher concentrations of NH3-N in the later stages of the experiment. As the major N component excreted by freshwater fishes, this suggests that CID combines strong feeding and excretory abilities [30]. Furthermore, submerged macrophytes play a crucial role in nitrogen cycling by absorbing NH3-N as their primary nitrogen source [31]. In the absence of sufficient macrophyte biomass due to heavy consumption by CID, the capacity of the system to utilize excess NH3-N is significantly reduced. This may partly explain the sustained increase in NH3-N concentrations in the later stages of the CID treatment.
Unlike P, CAU did not exert a significant effect on N fractions during the experiment. In CAU, P fractions were dominated by PP, and ISS showed a significant positive correlation with TP, confirming that the key mechanism by which CAU affects P content in water is the re-suspension of sediment due to the disturbance effect, which is consistent with the results of the previous study on benthic fishes [32]. In addition, the adsorption of dissolved forms of P in water onto the suspended matter and their settling or conversion to particulate P may be one of the reasons for the lack of significant changes in the SRP of CAU.
N fractions in water of the three treatments increased significantly throughout the experiment (two-way ANOVA, p < 0.05), whereas SRP did not differ significantly either in time or between treatments, which may be attributed to three aspects. First, the initial growth of H. verticillata in the three treatments required the uptake of a large amount of nutrients. Thus, NH3-N, as the main N source, decreased from day 0 to day 7, which may promote sediment nutrient release. Second, H. verticillata was so little affected by CAU that their biomass was not significantly different from NF. Thus, the subsequent elevation of N fractions in CAU and NF may have originated from the nutrients released by the yellowing and rotting of the lower leaves due to the lack of light, while CID dominates ingestion, digestion, and excretion. Third, the molar ratios of N:P during the experiment were greater than the Redfield ratio of 16:1 in all three treatments [33], exhibiting P limitation. In addition, the nutrient concentrations remained low concerning the limitation thresholds for phytoplankton proliferation (approximately 0.1 mg/L for TP and 1.2 mg/L for TN) [34]. Therefore, there were no significant changes in SRP during NH3-N continuously increased.

4.3. Impact of Fish on Phytoplankton

The result of initially higher Chl. a in CAU suggests that light attenuation due to sediment resuspension was insufficient to counteract the effect of increased nutrients on phytoplankton growth. Meanwhile, CAU also feeds on a portion of zooplankton, which promotes algal growth. Our results agree with the previous finding on benthic fish Acheilognathus macropterus [35]. With the growth of H. verticillata, the production of allelopathic substances increased accordingly, and then the growth of plankton was suppressed [36]. On the contrary, H. verticillata was consumed in CID and the inhibition of phytoplankton growth was reduced, resulting in phytoplankton in CID not being significantly different from that in CAU at the later stage. Since cyanophytes are usually smaller and lighter in weight, differences in relative abundance and relative biomass appeared in CAU and NF. Cryptophytes proliferated rapidly in CID due to a preference for nitrogen-rich waters and thus were significantly higher than NF at the end [37].

4.4. Impact of Fish on Sediment

The significant positive correlation between OM and TP in sediments suggests that P adsorption in sediment is related to OM, which can increase phosphate adsorption by influencing mineral (iron and aluminum) crystallinity or forming ternary complexes [38]. However, the alteration of OM in CID caused by heavy defecation did not promote the content of TP and NaOH-P, which could be ascribed to the humus formed by feces at a later stage that competes with P for adsorption sites or wraps mineral surfaces to inhibit P adsorption [39].
The disturbance behaviors, such as swimming and foraging, of the fish can promote gas exchange at the sediment interface [40]. At the same time, the radial oxygen loss of H. verticillata can increase the redox potential of sediment, which can promote the formation of iron curtain (iron oxides) and facilitate the sequestration of upward-diffusing P [41]. These two factors may together result in a significant increase in NaOH-P content in CAU. The significant elevation of HCl-P in CAU was caused by the rapid growth of H. verticillata. During photosynthesis, they absorb carbon dioxide or bicarbonate ions. This leads to an increase in the water pH, resulting in the oversaturation of CaCO3 and the formation of CaCO3-P co-precipitation, i.e., HCl-P [42].

4.5. Adsorption and Release of P from the Sediment

Sediment EPC0 often evaluates the direction and strength of P adsorption/desorption, and the higher values indicate the greater potential for P desorption [43]; thus, the EPC0 and ERI showed a consistent trend in magnitude among the three treatments. The shorter digestive tract of CID may cause the relatively low digestibility of H. verticillata. Fecal OM can alter the P release characteristics of sediment by altering dissolved organic matter components, sediment particle size, and functional groups [44]. The highest EPC0 and ERI in CAU can be explained by two reasons. From the sediment perspective, there was a significant increase in the NaOH-P content as a potentially mobile P. While, in overlying water, cyanophytes, as the dominant phytoplankton species in the later CAU, have a high affinity for P and luxury consumption of P, which provide the potential to pump P from sediments [45,46]. Moreover, the significant positive correlations between Chl. a, TP, phytoplankton biomass, and the ERI further suggest that phytoplankton can indirectly affect sediment P release through the utilization of P in water, which was also found in the previous study [47].

4.6. P Cycling Affected by Fish

The influence of CID and CAU on nutrients in water has two main pathways. On the one hand, the disturbance and excretory of fish can increase nutrients in water. On the other hand, the increased nutrients in water were taken up by phytoplankton, creating a negative interaction between the two latent variables. In addition, there are three pathways to influence sediment P release. First, water properties can indirectly affect sediment P release through the sediment–water interface environment, e.g., changes in ORP in water can affect the decomposition of NaOH-P [48]. Second, the accumulation of potentially bioavailable P (NaOH-P) in sediment directly contributes to an elevated ERI. Third, the direct positive effect of sediment properties is mainly influenced by its pH, with high pH facilitating the release of NaOH-P and low pH facilitating the release of HCl-P [49]. Feces in CID and the H. verticillata roots in CAU and NF secreted organic acids are considered to be the main cause of sediment P desorption due to the sediment pH alteration [50].

5. Conclusions

The results from this 28-day mesocosm experiment showed that CID slowed down the growth of H. verticillata via feeding while CAU had no significant effect. During the experiment, CID showed insufficient effect on nutrient levels in water but CAU elevated the levels of TSS, ISS, TP, and PP due to the disturbance of the sediment. The disturbance of CAU also significantly increased the fraction of NaOH-P and HCl-P in the sediment. Both CID and CAU elevated the ERI by altering the properties of water and sediment, as well as the P fractions of sediment, with CAU having a greater impact. Our preliminary results indicate that typical herbivorous fish decrease the growth of submerged macrophytes without affecting water quality while omni-benthivorous fish cause particle suspension without affecting the growth of submerged macrophytes. In the future, the influence of fish size and density and different types of submerged macrophytes should be studied to better guide fish management in shallow lakes.

Supplementary Materials

The following supporting information can be downloaded at https://www.mdpi.com/article/10.3390/w17030326/s1, Text S1: The equilibrium P concentration (EPC0); Figure S1: Schematic diagram of the location and experimental design of the Liangzi Lake field research station; Figure S2 The modified SMT protocol for different fractions of P. OP was obtained by subtracting IP from TP. All extracts were centrifuged, the supernatants were filtered through 0.45 μm GF/C filter membrane, and the phosphorus concentration was determined using the molybdenum blue methods; Figure S3 Correlation analysis among parameters in the overlying water and sediment; Table. S1 Differences among treatments during the experiment based on Two-way ANOVA; Table S2 The direct and indirect relationships between variables. The path coefficients are calculated by the partial least squares path model (PLS-PM).

Author Contributions

Conceptualization, C.W.; Methodology, S.H.; Formal analysis, W.W.; Investigation, X.C. (Xin Chen), W.W. and H.A.; Resources, H.A.; Data curation, X.C. (Xin Chen); Writing—original draft, X.C. (Xin Chen); Writing—review & editing, C.W.; Supervision, X.C. (Xiaofei Chen) and C.W.; Project administration, S.H.; Funding acquisition, H.C. and X.C. (Xiaofei Chen). All authors have read and agreed to the published version of the manuscript.

Funding

This research was funded by the Local Cooperation Project of China Academy of Engineering (HB2023B05) and the Ecological and Environmental Protection Research Project of Hubei Province (2022HB07).

Data Availability Statement

Data available on request due to restrictions eg privacy or ethical.

Conflicts of Interest

Author Shenghua Hu and Huaqiang Chen were employed by Wuhan Bridge Engineering Co., Ltd. The remaining authors declare that the research was conducted in the absence of any commercial or financial relationships that could be construed as a potential conflict of interest.

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Water 17 00326 g001
Figure 1. The mean biomass of fish (A) and H. verticillata (B) in all treatment groups at the beginning and the end of the experiment. Lowercase letters (a, b) above the bars indicate significant differences among treatments. Treatments sharing the same letter are not significantly different.
Figure 1. The mean biomass of fish (A) and H. verticillata (B) in all treatment groups at the beginning and the end of the experiment. Lowercase letters (a, b) above the bars indicate significant differences among treatments. Treatments sharing the same letter are not significantly different.
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Figure 2. Changes in pH (A), DO (B), and ORP (C) in overlying water during the experiment. Lowercase letters (a, b, c) above the bars indicate significant differences among treatments. Treatments sharing the same letter are not significantly different.
Figure 2. Changes in pH (A), DO (B), and ORP (C) in overlying water during the experiment. Lowercase letters (a, b, c) above the bars indicate significant differences among treatments. Treatments sharing the same letter are not significantly different.
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Figure 3. Changes in TSS (A), ISS (B), TP (C), PP (D), SRP (E), TN (F), TDN (G), NO3-N (H), and NH3-N (I) in overlying water during the experiment. Lowercase letters (a, b) above the bars indicate significant differences among treatments. Treatments sharing the same letter are not significantly different.
Figure 3. Changes in TSS (A), ISS (B), TP (C), PP (D), SRP (E), TN (F), TDN (G), NO3-N (H), and NH3-N (I) in overlying water during the experiment. Lowercase letters (a, b) above the bars indicate significant differences among treatments. Treatments sharing the same letter are not significantly different.
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Figure 4. Chl. a (A) relative and total abundance (B) and biomass (C) of phytoplankton measured from different treatments during the experiment. Lowercase letters above the bars indicate significant differences among treatments. Treatments sharing the same letter are not significantly different.
Figure 4. Chl. a (A) relative and total abundance (B) and biomass (C) of phytoplankton measured from different treatments during the experiment. Lowercase letters above the bars indicate significant differences among treatments. Treatments sharing the same letter are not significantly different.
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Figure 5. Changes in pH (A), OM (B), TP (C), Ex-P (D), NaOH-P (E), HCl-P (F), IP (G), and OP (H) of sediment during the experiment. * indicate levels of statistical significance: * p < 0.05, ** p < 0.01.
Figure 5. Changes in pH (A), OM (B), TP (C), Ex-P (D), NaOH-P (E), HCl-P (F), IP (G), and OP (H) of sediment during the experiment. * indicate levels of statistical significance: * p < 0.05, ** p < 0.01.
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Figure 6. Changes in EPC0 (A) and ERI (B) of sediment during the experiment. Lowercase letters (a, b, c) above the bars indicate significant differences among treatments. Treatments sharing the same letter are not significantly different.
Figure 6. Changes in EPC0 (A) and ERI (B) of sediment during the experiment. Lowercase letters (a, b, c) above the bars indicate significant differences among treatments. Treatments sharing the same letter are not significantly different.
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Figure 7. PLS-PM model analysis for P cycling. Dashed and solid boxes represent measured and latent variables, respectively. Solid lines indicate significant paths (* p < 0.05, ** p < 0.01), and the dotted lines indicate no significant paths. Red and blue lines indicate a positive and negative path, respectively. The thickness of the line indicates the magnitude of the path coefficient. The number on the red or blue line indicates the path coefficient.
Figure 7. PLS-PM model analysis for P cycling. Dashed and solid boxes represent measured and latent variables, respectively. Solid lines indicate significant paths (* p < 0.05, ** p < 0.01), and the dotted lines indicate no significant paths. Red and blue lines indicate a positive and negative path, respectively. The thickness of the line indicates the magnitude of the path coefficient. The number on the red or blue line indicates the path coefficient.
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MDPI and ACS Style

Chen, X.; Wu, W.; Ao, H.; Hu, S.; Chen, H.; Chen, X.; Wu, C. Impact of Grass Carp and Crucian Carp on Submerged Macrophyte and Phosphorus Cycling in Shallow Lake Mesocosms. Water 2025, 17, 326. https://doi.org/10.3390/w17030326

AMA Style

Chen X, Wu W, Ao H, Hu S, Chen H, Chen X, Wu C. Impact of Grass Carp and Crucian Carp on Submerged Macrophyte and Phosphorus Cycling in Shallow Lake Mesocosms. Water. 2025; 17(3):326. https://doi.org/10.3390/w17030326

Chicago/Turabian Style

Chen, Xin, Weiju Wu, Hongyi Ao, Shenghua Hu, Huaqiang Chen, Xiaofei Chen, and Chenxi Wu. 2025. "Impact of Grass Carp and Crucian Carp on Submerged Macrophyte and Phosphorus Cycling in Shallow Lake Mesocosms" Water 17, no. 3: 326. https://doi.org/10.3390/w17030326

APA Style

Chen, X., Wu, W., Ao, H., Hu, S., Chen, H., Chen, X., & Wu, C. (2025). Impact of Grass Carp and Crucian Carp on Submerged Macrophyte and Phosphorus Cycling in Shallow Lake Mesocosms. Water, 17(3), 326. https://doi.org/10.3390/w17030326

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