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Article

The Effect of the Potamogeton crispus on Phosphorus Changes throughout Growth and Decomposition: A Comparison of Indoor and Outdoor Studies

by
Lizhi Wang
1,*,†,
Liying Zhang
1,†,
Hongli Song
1,*,
Bin Dong
1,
Yun Wang
1,
Wanni Yu
1,
Yuanzhi Wu
1,
Xiaodong Wu
2 and
Xuguang Ge
2
1
Shandong Provincial Key Laboratory of Water and Soil Conservation and Environmental Protection, College of Resources and Environment, Linyi University, Linyi 276005, China
2
College of Urban and Environmental Sciences, Hubei Normal University, Huangshi 435002, China
*
Authors to whom correspondence should be addressed.
†
These authors contributed equally to this work.
Sustainability 2023, 15(4), 3372; https://0-doi-org.brum.beds.ac.uk/10.3390/su15043372
Submission received: 4 January 2023 / Revised: 7 February 2023 / Accepted: 10 February 2023 / Published: 12 February 2023
(This article belongs to the Special Issue Wetlands: Conservation, Management, Restoration and Policy)
Browse Figures
Versions Notes

Abstract

:
Phosphorus (P) transport and transformation in water were investigated using Potamogeton crispus. To compare and evaluate our indoor experiment with outdoor data, we used the simultaneous indoor experiment and field observation approach. The effects of P. crispus growth and decomposition on P concentrations were investigated. P. crispus significantly reduced the P content of different forms in the water during the growth period, and significantly increased the P content of different forms in the water during the decomposition period, according to the findings. As a result, the P level of the water varied seasonally and regularly. The pH and dissolved oxygen (DO) of environmental factors in the water revealed an increasing trend during the P. crispus growth period and a negative trend during the decomposition period. The changing trend of chlorophyll a (Chl-a) and alkaline phosphatase activity (APA) was inverse, decreasing during the growth period of P. crispus and increasing during the decomposition period. In the P. crispus growth environment, all forms of P in water were positively related to Chl-a, APA, and pH, and negatively related to DO. The comparison of the indoor experiment with field data revealed that the indoor experiment number has a larger standard deviation, indicating that the indoor experiment data fluctuated substantially. The indoor simulation experiment has the disadvantage of large data fluctuation. As a result, this study demonstrated that P. crispus regulated the P cycle in water via absorption and changes in environmental factors during the growth period, and released nutrients via decomposition during the decomposition period, thereby influencing the migration and transformation of P in the water. This work may be used as a reference for future research into the process of P exchange between sediments and water interfaces caused by P. crispus.

1. Introduction

Currently, more than 80% of China’s lakes are eutrophic, which is a significant sign of major water contamination. Eutrophication is primarily brought on by phosphorus (P), which increases enhanced positive feedback [1]. In lake ecosystems, submerged macrophytes play a key role in a variety of essential ecological processes, such as water filtration and environmental heterogeneity. They also provide abundant habitat and food supplies for fish, benthic invertebrates, and algae. In shallow lakes especially, submerged macrophytes play a significant part in the cycling of P [2]. Regarding the impact of submerged macrophytes on P cycling in lakes, however, there is still much to learn. P concentrations in lake water may rise or fall as a consequence of interactions between aquatic plants, sediment, and water [3].
Two alternative explanations are offered for these changes. In some cases, researchers reported that submerged plant communities were a net sink for nutrients [4]. By contrast, the alternative opinion is that the rooted macrophytes mediate a net transport of nutrients from the sediment to the water [5]. Some submerged macrophytes acquire significant amounts of nutrients through both shoots and roots from the water and the substrates, respectively; however, under normal pore and lake water P concentrations, the uptake of nutrients from the substrate is the primary process [6]. The growth of submerged macrophytes has minimal effect on the release of P from the sediment. Some submerged macrophytes (including P. crispus) have degenerated roots and obtain P primarily from the water, and submerged macrophytes with developed roots (such as Vallisneria natans) obtain P primarily from the substrate. Decaying submerged macrophytes may provide an internal source of P for the lake and release considerable quantities of P into the water [7]. Simultaneously, the sediment retains a large portion of the released P.
Submerged macrophytes are the key producers of shallow lakes, occupying the main interface of water and sediment in lakes, and regulating the material, energy, and transfer cycles of lake ecosystems [8]. During the growth period, plants can absorb P and regulate its transport and transformation in sediments and overlying water. As a result, converting eutrophic lakes to healthy grass-type lakes usually necessitates the reconstruction of aquatic vegetation dominated by submerged macrophytes.
Submerged macrophytes play a significant role in shallow lakes [2]. During growth, submerged macrophytes absorb nutrients in the water, and with the vegetation dieoff, the large biomass of these plants, particularly in eutrophic lakes, releases the nutrients back into the water [9]. Three mechanisms are associated with submerged macrophytes and the sequestering of P during plant growth [10]. First, the majority of the submerged macrophyte species assimilate P directly through the thalli, shoots, and leaves, which is a process that bypasses the soil nutrient reservoir. Second, the periphyton that grows on the submerged macrophytes also removes nutrients directly from the water. Third, the elevation in pH that results from the intense photosynthesis of the submerged macrophytes and periphyton leads to CaCO3 supersaturation, which may facilitate the removal of P via coprecipitation. Because of its low cost, environmental friendliness, and long-term effect, aquatic vegetation restoration has increasingly become a research hotspot of ecological remediation technology [11].
Throughout China, Potamogeton crispus is a submerged macrophyte that is widely used in the ecological rehabilitation of eutrophic lakes [12]. The effectiveness of aquatic vegetation has been credited with the majority of restoration successes in eutrophic lakes [13]. P. crispus germinates in the autumn and grows over the winter. In the spring and summer, the plants ultimately degrade and die. It expands quickly, can survive in conditions with many nutrients, and thrives in polluted water [14].
P. crispus, a submerged macrophyte, secretes allelochemicals that hinder the growth of algae, absorb nutrients from the water and sediment above them, and have significant ecological significance in reducing lake eutrophication [15]. Shallow lakes’ biological management and endogenous pollution load reduction have both benefited from the restoration, reconstruction, and alteration of aquatic vegetation [16]. However, when several submerged macrophytes decay, the leftover material remains in the water and decomposes, releasing nitrogen (N), phosphorus (P), and other unprocessed elements into the surrounding water causing secondary pollution [17]. Studying the P cycle of water throughout P. crispus’ whole life cycle is crucial, for this reason.
In light of the aforementioned issues, two study approaches are now in use: field observation and indoor simulation [18]. Because indoor simulation data are limited by constraints, it is frequently impossible to accurately reproduce field observations [19]. As a result, it is critical to investigate the link between indoor simulation experiments and field observations. Indoor simulation experiments are performed in this study to examine the differences between field data and indoor simulation data. It is envisaged that their interaction would be established as a reference for future research on water ecological restoration.
Lake Yimeng is a lake with submerged macrophyte P. crispus as the dominant species. We detect changes in P of the P. crispus’ entire life cycle in water using a combination of indoor studies and field observations. We also analyze the link between P transport and transformation in water. At the same time, the role of the environment in P transport and transformation is examined.
Hence, this study aimed (1) to compare and analyze the influence of P. crispus on P migration and transformation in water in field observation and indoor experiments, and (2) to analyze the relationship between P migration and transformation and environmental factors in water.

2. Materials and Methods

2.1. Study Area and Sampling

Lake Yimeng was produced in 1997 by the construction of a rubber dam (1135 m) over the River Yi trapping 12 million m3 of water. P. crispus has formed dense canopy-forming populations in the lake in recent decades, covering about 90% of the lake during spring and summer [20]. The River Beng section of the lake (P. crispus region) is connected to the River Yi section of the lake (35°2′40″–35°7′11″ N; 118°22’11″–118°23′12″ E). Two matching sample locations in the P. crispus and non-aquatic plant habitats were chosen (Figure 1).
During the growth season, P. crispus entirely covers the River Beng, with a mean biomass of 1025 g·m−2. Although the River Beng and the River Yi are linked, there is minimal water exchange. Water flows from the River Beng to the River Yi during the rainy season, which lasts from July to August. The Secchi disc depth in the River Yi ranges from 20 cm to 60 cm, and from 50 cm to 145 cm in the River Beng.
Every month from 4 January 2017 to 4 December 2019, mixed water samples were taken from the top layer of each sampling site (0, 0.5, and 1 m) using a Grasp sampler (Grasp BC-2300, Beijing, China). Following that, water samples were kept in a refrigerator at 4 °C for subsequent examination. Two liters of water were collected to test physical and chemical indicators such as chlorophyll a (Chl-a) and alkaline phosphatase activity (APA).

2.2. Indoor Experiments

2.2.1. Sediment

The surface sediment samples (the top 20 cm of sediment) were collected from Lake Yimeng. The samples were collected using a core sampler with a 50 cm LX, 10 cm DI Plexiglas cylinder tube in July 2018. They were transported to the laboratory in sealed plastic bags on ice. Upon arrival in the laboratory, the samples were freeze-dried, passed through a 0.5 cm sieve to remove coarse debris, and mixed.
The homogenized sediment was added to high-density polyethylene buckets (top diameter = 55 cm, bottom diameter = 45 cm, and height = 75 cm). Each polyethylene bucket contained 4832.80 g of dry sediment. The depth of the sediment was 10 cm. A total of 100 L of deionized water was slowly added to the buckets. The samples were equilibrated for one month, and the sediment was collected. Preliminary tests were conducted on the water and the sediment at the beginning of the experiment (Table 1).

2.2.2. Experimental Submerged Macrophyte

The plant propagules of the submerged macrophyte P. crispus were collected from Lake Yimeng in May 2018 for experimental use. All plant propagules were centrifuged for approximately 15 s, brushed carefully to remove adherent water and materials, and weighed to determine the initial fresh weights (40 g). After the weights were recorded, the plant propagules were assigned at random for planting in three buckets that contained sediment, with 40 g of propagules planted in each bucket (i.e., the treatment condition). Additionally, three buckets were left unplanted to serve as the controls. All the buckets were placed in a greenhouse with ambient daylight. Among the buckets, the water temperature differed by less than 2 °C. The losses in culture volume from evapotranspiration were countered with the addition of deionized water to the original volume each month.

2.3. Sampling and Analyses

The indoor studies lasted a year, from 4 November 2018 to 31 October 2019. Each month, water was collected from 0.25 m under the surface and filtered over a 0.45 m cellulose acetate membrane. The concentrations of TP, particulate P (PP), dissolved total P (DTP), dissolved organic P (DOP), and soluble reactive P (SRP) in the water samples were determined using standard methods [21]. Before analysis, the samples were kept in a refrigerator at 4 °C. Beijing N&D Co., Ltd. (18 Jiaomen, Fengtai District, Beijing)in China supplied the cellulose acetate membranes.
Water samples of 50 mL were taken for TP, PP, DTP, DOP, and SRP analysis. Every 30 days, a suitable volume of deionized water was introduced to additional beakers to compensate for water loss due to evaporation. The P concentration was calculated using the following equation:
C n = c n ′ v v + v n
where Cn is the actual P concentration at the n-th sampling time, c n ′ is the measured P concentration before adding deionized water, v is the volume of overlying water before adding deionized water, and vn is the volume of deionized water added at time n.
Because the plant biomass could not be directly assessed throughout the experiment, the biomass was calculated using a model. During the growth phase, the length (X/cm) and biomass (wet weight, W/g) of P. crispus were measured, and the length included the branches of the submerged macrophyte. A single linear regression model was used to calculate the biomass (W) of P. crispus:
W = a X + b
where W is the biomass of the submerged macrophyte, X is the length of the submerged macrophyte, and a and b are constants. The model to estimate the biomass for P. crispus was as follows:
W P. crispus = 0.0487X − 1.1699; R2 = 0.76; p < 0.05

2.4. Determination of Physical and Chemical Indexes

The water sample was autoclaved at 121 °C for 30 min following the addition of K2S2O8 for the TP analysis. The sample was then treated with 10% ascorbic acid and analyzed using the molybdenum blue spectrophotometric technique. The SRP was determined using an identical procedure, with the exception that the water sample was filtered through a 0.45 μm cellulose acetate membrane and was not autoclaved. The DTP was determined in the same way as the TP, but the water was filtered through a 0.45 μm cellulose acetate membrane. The PP form was defined as the difference between the TP and the DTP. The DOP form was defined as the difference between the DTP and the SRP. SRP and DTP concentrations in water have a detection limit of 1 μg·L−1. The amounts of P were determined using a continuous flow analyzer (FLOWSYS III Italy Systea business). Shanghai N&D Co., Ltd. provided all of the materials for these studies (China). All samples were examined in triplicate, and the results are given as the mean.
The pH (PHSJ-4A, Lei-ci, Shanghai, China) and the DO (YSI 5750, Columbus, OH, USA) in the water were measured before the collection of the samples. The sediment samples were collected using a hand-driven stainless steel corer, and were transported to the lab immediately in sealed plastic bags in iceboxes; in the laboratory, the sediment was freeze-dried and sieved (<2 mm). The fractionation procedure was based primarily on the SMT protocols [22]. The sediment samples that were analyzed for the Olsen-P were extracted from 1 g of sediment with 20 mL of 0.5 M NaHCO3 (pH = 8.5) for 30 min [23]. The organic matter content was determined after treatment of the sample with K2Cr2O7/H2SO4 according to the Walkey–Black method [24]. The cation exchange capacity was analyzed using the EDTA-NH4+ method, and the TN was measured using the concentrated H2SO4 digestion method [25].
The extracellular APA in the water samples was measured using Gage and Gorham’s [26] previously reported procedures. To evaluate the Chl-a content, water samples were filtered through 0.45 m Millipore filters and extracted with 90% acetone [27].

2.5. Statistical Analysis

SPSS v 25.0 was used for correlation analysis (IBM Corp., Chicago, IL, USA). p < 0.05 and p < 0.01 were used to determine statistical significance. Tukey’s post-hoc test was utilized after a two-way ANOVA to identify the differences between the indoor experiment and the outdoor investigation. Pearson correlation analyses were used to analyze the link between the chemical and physical properties of the water in the research locations. The average of three replicates were used to calculate the findings. The maps were created using ArcMap 10.8 (https://www.esri.com/ (accessed on 1 March 2021)).

3. Results

3.1. Physical and Chemical Indexes of Lake Yimeng

The TP concentration in the two bodies of water was the lowest in winter and the highest in summer (Figure 2a). TP rapidly increased from April to June and peaked in summer (from May to August). For three years, TP in the P. crispus area remained low (0.04–0.06 mgL−1) from January to March. The TP in the non-aquatic plant area showed the same seasonal variation tendency. TP in the non-aquatic plant area was significantly lower than that of P. crispus (p < 0.05). The TP gradually increased from May to October, reaching its second high in October. The TP trend of the non-aquatic plant area varied in the same way as the P. crispus area from 2018 to 2019, however, the TP concentration was comparatively low in summer and high in winter.
The DTP in the P. crispus area had a similar trend to TP and peaked in summer (from May to August). DTP showed an increasing trend in the area of non-aquatic plants, with the highest concentration being discovered in August 2019. In comparison to the P. crispus area, the DTP in the non-aquatic plant area was significantly lower (p < 0.05). Compared to the P. crispus area, DTP gradually increased and plateaued from winter to summer (Figure 2b). In neither water body was the changing trend in PP evident, but it appeared to increase in summer and decrease in winter (Figure 2c). According to data from various years, there was no significant difference between the non-aquatic plant and P. crispus areas of PP in 2017–2018 (p > 0.05), but in 2019 the content of PP in the non–aquatic plant area increased significantly. SRP increased significantly in the P. crispus area from May to July of 2017, and it increased once more from May to June in 2018 and 2019. After reaching its peak, the SRP gradually dropped (Figure 2d). In general, from 2017 to 2019, TP, DTP, and SRP showed rising trends in the non-aquatic plant area but declining trends in the P. crispus area.
DOP was kept at a high level (0.05–0.14 mg·L−1) and increased from May to September in the P. crispus area. When compared to the non-aquatic plant area, it was significantly higher in the P. crispus area (p < 0.05). DOP was kept at a low level and only slightly altered in the area of non-aquatic plants (Figure 2e). In the P. crispus region, TP, DTP, PP, SRP, and DOP increased and were consistent with P. crispus’s senescence period.
The Chl-a content change followed the same pattern as the TP, DTP, and SRP changes. In the P. crispus and non-aquatic plant areas, the Chl-a content had a positive relationship with the TP, DTP, and SRP concentrations (Figure 2f). Additionally, in 2017 there was a significant difference in Chl-a content between the P. crispus and non-aquatic plant areas (p < 0.05). During the P. crispus growth period from 2018 to 2019, the non-aquatic plant area’s Chl-a content was significantly higher than that in the P. crispus area. On a seasonal scale, winter had the lowest Chl-a content, and summer had the highest. In April, the Chl-a content sharply increased when P. crispus began to decompose.
The pH and DO in the P. crispus region rose throughout the growth phase from March to May and declined during the decomposition period. In the non-aquatic plant area, the pH was lowest in winter and greatest in summer, but the DO seemed to follow the reverse trend (Figure 2g,h).
Every year, from March to May, P. crispus experienced a rapid growth phase during which time its biomass in the water body rapidly rose. After June, P. crispus entered its decline phase, and as a result of decay and decomposition, its biomass gradually decreased in the water body. P. crispus biomass in the water body thus displayed a single peak change trend (Figure 2i).
From July to September, the total APA in the P. crispus region remained high. The APA of the non-aquatic plant area did not vary significantly from the APA of the P. crispus area (p > 0.05). From April to December, the overall APA remained reasonably high. The total APA in the two bodies of water was lowest in winter and largest in summer (Figure 2j).

3.2. Physical and Chemical Indexes of Water in the Indoor Experiment

The treatment group’s TP concentration fluctuated more than the control group’s (Figure 3a). When compared to the control group, the TP in the treatment group declined dramatically from November 2018 to April 2019, then rose significantly from April 2019 to July 2019. From August 2019 to October 2019, the TP concentration stabilized, and no significant differences between the control and treatment groups were observed. The treatment group had the greatest TP concentration (0.57 mg·L−1) in June 2019, and the lowest TP concentration (0.017 mg·L−1) in January 2019. The greatest and lowest amounts of TP in the control group’s water were 0.21 and 0.05 mg·L−1, respectively, and were detected in February and April of 2019.
The fluctuations in the concentrations of the DTP and TP were similar in the treatment group. The concentration of the DTP in the water was not significantly different between the control and treatment groups from November 2019 to April 2019 or from August 2019 to October 2019. However, the concentration of the DTP increased significantly in the treatment group from April 2019 to July 2019 (Figure 3b). The concentrations of the DTP in the treatment group increased from 0.04 to 0.45 mg·L−1 from April 2019 to June 2019, but decreased rapidly to 0.11 mg·L−1 in August 2019. The highest concentration of the DTP (0.45 mg·L−1) was in the treatment group in June 2019, and the lowest value of the DTP (0.016 mg·L−1) was in January 2019. The highest and the lowest concentrations of the DTP in the control group were 0.08 and 0.04 mg·L−1, respectively, which were recorded in February and April 2019, respectively.
Compared to the control group, the PP decreased significantly in the treatment group from November 2019 to April 2019 but increased significantly in the treatment group from April 2019 to June 2019. No significant differences were detected between the control and the treatment groups from September 2019 to October 2019 (Figure 3c). From the beginning of the experiment, the concentrations of PP in the treatment group decreased rapidly to 0.002 mg·L−1 in January 2019. From January to April 2019, the PP remained at a low level and ranged from 0.002 to 0.042 mg·L−1; however, simultaneously in the controls, the concentrations of PP were at a high level and ranged from 0.08 to 0.13 mg·L−1. The highest value of the PP (0.12 mg·L−1) was in June 2019.
Compared to the control group, the concentrations of SRP decreased significantly from November 2019 to April 2019 in the treatment group. However, they increased significantly in the treatment group from April 2019 to October 2019 (Figure 3d). The SRP remained at a low level (from 0.08 to 0.06 mg·L−1) in the water from November 2019 to April 2019. SRP concentrations of the control group remained at low levels during the experiment.
For the DOP, no significant differences were found in the water between the control and the treatment groups from November 2019 to April 2019. The concentration of DOP in the treatment group increased rapidly to 0.16 mg·L−1 from April 2019 to May 2019 and then decreased rapidly to 0.06 mg·L−1 in July 2019 (Figure 3e). The highest DOP value in the treatment group (0.16 mg·L−1) occurred in May 2019, but the highest values of the other P forms occurred in June 2019. The concentrations of DOP in the control group remained at a low level and ranged from 0.002 to 0.05 mg·L−1.
During the P. crispus growth period from November 2018 to May 2019, the control’s Chl-a content was significantly higher than that in the P. crispus group (p < 0.05). The Chl-a content increased during the growth decomposition period of P. crispus from June 2019 to July 2019 (Figure 3f).
The pH and DO in the P. crispus group rose throughout the growth phase from November 2018 to May 2019 and declined during the decomposition period (Figure 3g,h). During the growing phase of P. crispus from November 2018 to May 2019, the pH and DO were significantly higher in the P. crispus group compared to the control group (p < 0.05).
From March to May 2019, P. crispus experienced a rapid growth phase during which time its biomass in the P. crispus group rapidly rose. After June 2019, P. crispus entered its decomposition phase, and as a result of decay and decomposition, its biomass gradually decreased (Figure 3i).
Throughout the experiment, the total APA in the P. crispus group was significantly higher than in the control group (p < 0.05). The total APA grew throughout the P. crispus growth phase from November 2018 to May 2019 and declined during the P. crispus decomposition period (Figure 3j).

3.3. The Pore Water Concentrations of the Indoor Experiment

The concentrations of all the P forms in the pore water were higher than those in the overlying water during the growth of P. crispus (Figure 4a–c). The P in the pore water tended to infiltrate into the overlying water. The concentrations of all the P forms in the water column increased rapidly during the decomposition of P. crispus. The equilibrium of P between the pore water and the overlying water was disrupted, and the P in the overlying water was absorbed by the sediment and infiltrated into the pore water. Thus, the concentrations of P in the pore water in the treatment group were higher than those in the control group. The complicated reciprocal actions between the P sorption and desorption on the sediment resulted in changes in the concentrations of P in the overlying water and pore water.

3.4. Comparison of Indoor and Field Data

The comparison of indoor and field data revealed that there was no significant difference between the forms of P in the indoor experimental water and those in the field, showing that indoor trials could imitate field settings. The standard deviation of each experimental index in the indoor experiment, on the other hand, was larger than that in the outdoor experiment, indicating that the indoor experimental data have a high level of variability. Simultaneously, there were substantial changes in certain environmental indicators between indoor and outdoor studies. Therefore, while the indoor experiment could imitate the condition of the field experiment, owing to the relatively tiny simulation system, many elements were difficult to regulate, resulting in rather dramatic fluctuations in many indicators of the water body (Table 2). This should be connected to the lesser buffer capacity of the indoor simulation system.
A comparison of the indoor experiment and field data indicated the indoor experiment still has many flaws and instabilities. At the same time, owing to the study’s detection settings’ limits, it was unable to identify the numerous indoor and outdoor indicators thoroughly, and the analysis was insufficient. The findings of this research may be used to compare indoor and outdoor trials.

4. Discussion

4.1. Effect of P. crispus on the Concentrations of P Forms in the Overlying Water in Lake Yimeng

In this field investigation case study, we compared the changes in different P concentrations triggered by submerged macrophyte P. crispus. Our findings indicated that the presence of P. crispus reduced P concentrations in the water column in terms of SRP and TP, whereas its decay raised DOP concentrations. Phytoplankton reacted appropriately to changes in P content. When the submerged macrophyte P. crispus degraded, a phytoplankton bloom formed because of P released from P. crispus.
The natural phytoplankton of the vegetated area (River Beng) was dominated by cyanobacteria, like the universal prediction that there was often a shift in the phytoplankton community towards dominance by cyanobacteria with high P concentrations. Cyanobacteria dominated the phytoplankton community in the studied lake, especially during the decomposition of P. crispus. This was in line with the indoor experiment of our study, revealing that further enrichment with N plus P in eutrophic water would enhance cyanobacterial dominance. So, the formation of P and the P-using strategy jointly shaped the pattern of the phytoplankton community.
Under indoor simulation circumstances, the release of P from the breakdown of submerged macrophytes typically takes around 30 days. Our findings concur with those of Wang et al. [28], who found that following decomposition, P. crispus emitted a significant quantity of DOP. During P. crispus’ development stage, different P concentrations in the River Beng section of the lake—where P. crispus was present—were considerably lower than those in the River Yi part of the lake, indicating that P. crispus’ presence reduced the amount of bio-available P in the water column. When P. crispus decomposed, the DOP content in the River Beng portion of the lake rose and exceeded that in the River Yi portion, indicating that some OP was released during the breakdown process of P. crispus. For the lake’s phytoplankton population throughout the summer and autumn, P. crispus decomposition represented a significant source of biologically accessible P and decomposable dissolved organic matter, indicating that P. crispus decomposition may drive phytoplankton production [29]. According to Stephen et al. [30], who disagreed with our findings, the submerged macrophyte P. crispus enhanced the rate at which P was released into the sediment throughout the growing season. During its development phase, P. crispus took P from the sediment, and during its breakdown phase, it released P back into the water. Increased P concentrations in the surrounding waters were brought about by P. crispus development, which also moved P from the sediment to the surrounding water [31].
The majority (about 47%) of the P in the water in the River Yi section was made up of PP. In the absence of P. crispus, the water body was significantly agitated by the wind and waves, raising the PP concentration. Unstable bound P on the surface of certain particles undergoes desorption, reduction, or dissociation events that transform it into SRP and other P fractions as a result of changes in environmental restrictions [32]. Therefore, eutrophication is more likely when SRP and PP concentrations are high.

4.2. Effect of P. crispus on the Concentrations of P Forms in the Overlying Water in the Indoor Experiment

According to our research, the submerged macrophyte P. crispus developed rapidly and increased the water’s ability to absorb P throughout the growth phase. The two optimum seasons for P. crispus growth were winter and spring. As a result, in the winter and spring, P. crispus can cause the TP levels in the neighboring streams to decrease. The TP concentration in the treatment group fell to a low level after two months of culture (January 2019), and then tended to level out during the next two months. The increase in TP content in the control group’s waters in February 2019 was most likely brought on by the P released from the sediment and the absorption by plants. This finding could have been caused by the macrophyte P. crispus, which was buried and delayed the release of P from the sediment, increasing P absorbability and causing P transformation in the water. Although the removal of P from the water was considerably aided by the submerged plants’ absorption of P, P loss from the overlying water system could not be completely explained, indicating that other biochemical and physicochemical processes may also be at play [33]. The P variations seen in the control group suggest that these activities produce a fall in P levels and an increase in P levels in the adjacent waters.
P. crispus developed vigorously throughout the growth phase due to the direct absorption of DOP and SRP from the surrounding water [34]. In the treatment group, DOP and SRP concentrations accordingly dropped and stayed at low levels throughout P. crispus development.
Due to the static settings of the experiment, DTP was the predominant form of P in the waters. Since P. crispus absorbed DOP and SRP from the water and controlled the P release from the sediment throughout the growth phase, DTP practically did not change during this time.
The changes in the various P forms are significantly influenced by PP [35]. The orthophosphate in PP may be transformed into dissolved P or absorbed by the sediment in addition to being directly absorbed and used by submerged macrophytes [36]. Due to the disruption caused by planting the P. crispus propagules and the consequent release of P from the sediment into the surrounding water, the concentration of PP was fairly high at the start of the experiment. During the development phase, P. crispus may also collect particles from the surrounding water. As a result, the PP in the water quickly dropped during the growth phase before being released again during the breakdown phase. PP was not the predominant P form in the surrounding water, although its concentration rose over the decomposition period.
Because P. crispus actively absorbed the P in pore water with both roots and shoots during growth, the concentration gradients between the pore water and the overlying water in the control group were greater than those in the treatment group. P concentrations in the pore water and the water above it dropped as a result of the treatment. The balance of P between the pore water and the overlying water was disturbed, which led to more P infiltrating the overlying water for absorption by P. crispus. This was another factor contributing to the reduction of P in the pore water.
The release of P might be one of the factors contributing to the rise in P content in the surrounding water in the treatment group between August and October 2019. As a result, the treatment group’s pore water P concentrations were lower than those of the control group. Another explanation is that P. crispus propagule development and absorption of P from the pore waters was the source of the reduction in P concentration in pore waters.
Based on the results of this study, the pore water was a sink for P when the concentrations of P in the overlying waters were higher than the equilibrium between the surrounding water and the pore water. When the concentrations of P in the overlying waters are lower than the equilibrium between the surrounding water and the pore water, the pore water is a resource for P. Thus, the pore water buffers the concentrations of P between the sediment and the overlying waters. This buffering ability improves in the treatment with P. crispus.

4.3. Effects of Environmental Factors on the Concentrations of P Forms in the Overlying Water

The environmental factors that influenced the release of P from lake sediment primarily included temperature, pH, redox potential, and hydrological conditions. Among these factors, DO and pH were the primary factors, and the effects of DO and pH on P release were investigated in many previous studies [37]. In the treatment group and the field P. crispus area, the concentration of DO and the pH increased in the growth period and decreased in the decomposition period, which was consistent with previous studies [38]. P. crispus used carbon dioxide in photosynthesis and released O2 under light conditions in the growth period. Moreover, the processes of photosynthesis, biological N fixation, and denitrification may cause an increase in the pH [39], and the photosynthetic process might be the primary reason that the pH and the concentration of DO increased in the surrounding waters. In the decomposition period, only some algae used carbon dioxide in photosynthesis, and the decomposition of P. crispus required a substantial quantity of oxygen. Therefore, the pH and the concentration of DO decreased rapidly. Such results have been observed in field observation and the indoor experiment.
In conditions with light, the photosynthetic process causes an increase in the concentration of oxygen. In aerobic conditions, the P binds to Fe3+ to form Fe2(PO4)3 [40]. Simultaneously, the Fe(OH)3 in the sediment absorbs the dissolved P in the overlying water. Therefore, in an aerobic system, the release of P from the sediment is difficult. In an anoxic and anaerobic system, the P is easily released from the sediment into the overlying water. Throughout the experiment, the concentration of oxygen remained high. Therefore, the P was absorbed into the sediment, and the ratio of the P adsorption to the sediment positively correlated with the oxygen concentration [41].
The concentration of different P concentrations in the surrounding water negatively correlated with the concentration of DO in the treatment group and the field P. crispus area (Figure 5), but the correlation was not significant. Thus, the high concentrations of DO indicated a high absorption ability of the sediment and provided another reason for the lower concentrations of all P forms in the treatment group than in the control group in the overlying water.
The pH in the overlying water was affected by levels of oxygen, and the pH increased with an increase in the concentration of DO (Figure 5), which was one of the most important mechanisms affecting the release of P [42]. The pH of the overlying water was affected by many factors, particularly biological factors. In the treatment group and the field P. crispus area, the pH of the overlying water increased with the photosynthesis of P. crispus. The pH then remained high in the decomposition of P. crispus during the photosynthesis of the algae. Biological factors played an important role in the increase in the pH of the overlying water under anaerobic conditions [43]. An increase in the pH reduced the availability of the binding sites on ferric complexes because of the competition between the hydroxyl ions and the iron-bound P [44]. Moreover, a low pH might result in the reduction of ferric to ferrous iron, which would inhibit the absorption of P [45]. Thus, the concentration of P in the overlying water negatively correlated with the pH in the treatment group and the field P. crispus area (Figure 5), providing another explanation for the decrease in the P forms in the overlying water during the growth of P. crispus. Furthermore, the pH might be the primary factor affecting the release of P at the sediment-water interface.
A P hydrolase with broad specificity alkaline phosphatase may catalyze the hydrolysis processes of all P esters as well as the transfer reactions of P groups. It actively participates in P metabolism, supplies P for the water body’s fast plankton development, and is crucial to the transformation of P in aquatic ecosystems [46]. In this research, APA rose in the summer and fell in the winter. Another potential explanation is that phytoplankton may overcome the P deficit by excreting extracellular phosphatase, which catalyzes DOP [47]. Alkaline phosphatase in the water was altered by P. crispus, which decreased the maximal rate of enzymatic reaction. An earlier investigation showed that the total APA in an Indian pond system was inversely related to the total TP and bacteria in the water [48]. To the treatment group and the field P. crispus region, APA was found to be correctly applied in this investigation (Figure 5). This study’s findings are in line with those of earlier investigations.
In river ecosystems, phytoplankton is a significant producer that also contributes significantly to the processes of material exchange and energy oxidation [49]. Chl-a is a comprehensive performance that represents biological, physical, and chemical aspects of the water body and is constrained by external constraints [50]. Chl-a may therefore be used to describe the development stage of phytoplankton in water. In the treatment group and the P. crispus field region, the Chl-a content significantly correlated positively with the TP, DTP, PP, SRP, and pH. (Figure 5). This showed that Chl-a concentration is significantly impacted by the development and decay of P. crispus in water.

5. Conclusions

The P forms in the overlying water fluctuated seasonally under the impact of P. crispus. P. crispus absorption reduced the concentrations of all P forms in the water column, which remained low throughout P. crispus growth. From April to July, P concentrations increased when P. crispus decomposed, and P. crispus decomposition largely released DOP into the underlying water.
The pore water served as a buffer between the sediment P concentrations and the above waters. P. crispus treatment improved this buffering capacity.
In the P. crispus growth environment, all forms of P in the water were favorably associated with chlorophyll a, APA, and pH, and negatively related to DO. There is a positive link between different types of P in the water. It was shown that environmental factors impacted the concentration of various forms of P in the water in a multifaceted manner.
The results from an indoor simulation experiment replicating the effect of P. crispus development on P in a water body showed a higher standard deviation than data from a field experiment. The P concentration of various forms in water bodies did not vary significantly between indoor simulation data and outdoor data. The results from the indoor simulation experiment changed substantially due to the system’s instability. The indoor simulation experiment has a significant flaw.

Author Contributions

All authors contributed to the conception and design of this study. Material preparation, data collection, and analysis were performed by L.W., L.Z. and H.S. The first draft of the manuscript was written by L.W. and W.Y., and data analysis was performed by B.D., Y.W. (Yuanzhi Wu) and Y.W. (Yun Wang). Y.W. (Yuanzhi Wu), X.W. and X.G. commented on the previous versions of the manuscript. All authors have read and agreed to the published version of the manuscript.

Funding

This work was supported by the National Natural Science Foundation of China (32071630 and 42277306) and Natural Science Foundation of Shandong Province, China (ZR2021MD045, ZR2021MD003, and ZR2020MD102). Innovation and Entrepreneurship Training Program for College Students in Shandong Province, China (202110452092).

Institutional Review Board Statement

Not applicable.

Informed Consent Statement

Not applicable.

Data Availability Statement

The data used to support the findings of this study are available from the corresponding author upon request (LZW [email protected] and HLS [email protected]).

Conflicts of Interest

The authors declare no conflict of interest.

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Figure 1. The location of study region and sample locations. ArcMap version 10.8 was used to create this map (https://www.esri.com/ (accessed on 1 March 2021)).
Figure 1. The location of study region and sample locations. ArcMap version 10.8 was used to create this map (https://www.esri.com/ (accessed on 1 March 2021)).
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Figure 2. Water chemical and physical parameter monthly fluctuations (mean ± SD) in several Lake Yimeng locations from January 2017 to December 2019. Varying lake locations have varied P fraction concentrations in the water (a–e), different Chl-a concentrations in the water (f), and different physical properties (g,h). (i) Biomass in various lake regions. (j) APA concentration in various lake locations.
Figure 2. Water chemical and physical parameter monthly fluctuations (mean ± SD) in several Lake Yimeng locations from January 2017 to December 2019. Varying lake locations have varied P fraction concentrations in the water (a–e), different Chl-a concentrations in the water (f), and different physical properties (g,h). (i) Biomass in various lake regions. (j) APA concentration in various lake locations.
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Figure 3. Water chemical and physical parameter monthly fluctuations (mean ± SD) under various treatments from November 2018 to October 2019. Physical parameters in various treatments include (a–e) water varied P fraction concentration, (f) water Chl-a concentration, and (g,h) physical parameters in the treatment, (i) biomass. (j) APA concentration in water under various conditions.
Figure 3. Water chemical and physical parameter monthly fluctuations (mean ± SD) under various treatments from November 2018 to October 2019. Physical parameters in various treatments include (a–e) water varied P fraction concentration, (f) water Chl-a concentration, and (g,h) physical parameters in the treatment, (i) biomass. (j) APA concentration in water under various conditions.
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Figure 4. Changes in the different forms of P in the pore water in the treatment and control groups (mean ± SD). (a) DTP in the pore water, (b) SRP in the pore water, (c) DOP in the pore water.
Figure 4. Changes in the different forms of P in the pore water in the treatment and control groups (mean ± SD). (a) DTP in the pore water, (b) SRP in the pore water, (c) DOP in the pore water.
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Figure 5. The correlations between water’s chemical and physical properties in the field study regions and the indoor experiment are described by Pearson’s correlation coefficients.
Figure 5. The correlations between water’s chemical and physical properties in the field study regions and the indoor experiment are described by Pearson’s correlation coefficients.
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Table
Table 1. Summary of the initial physicochemical parameters of the water and the sediment.
Table 1. Summary of the initial physicochemical parameters of the water and the sediment.
Sediment and WaterIndicatorsInitContent
TP(mg·kg−1)630.46
IP(mg·kg−1)497.01
OP(mg·kg−1)111.47
SedimentOlsen-P(mg·kg−1)75.98
TN(mg·kg−1)1810.13
OM(mg·kg−1)42,844.44
CEC(cmol·kg−1)160.32
TP(mg·L−1)0.10 ± 0.01
PP(mg·L−1)0.07 ± 0.01
DTP(mg·L−1)0.02 ± 0.00
SRP(mg·L−1)0.01 ± 0.00
WaterDOP(mg·L−1)0.01 ± 0.00
TN(mg·L−1)1.11 ± 0.12
N H 4 + –N(mg·L−1)0.19 ± 0.02
N O 3 − –N(mg·L−1)0.09 ± 0.00
DO(mg·L−1)9.57 ± 0.51
pH 7.62 ± 0.22
The values are the means ± standard deviation (n = 6), and the abbreviations indicate the following: TP = total P; IP = inorganic P; OP = organic P; Olsen-P = available P; TN = total N; OM = organic matter, CEC = cation exchange capacity; PP = particulate P; DTP = dissolved total P; SRP = soluble reactive P; DOP = dissolved organic P; N H 4 + –N = ammonium N; N O 3 − -N = nitrate N; DO = dissolved oxygen.
Table
Table 2. Comparative analysis of the indoor experiment and field data.
Table 2. Comparative analysis of the indoor experiment and field data.
ItemsExperiment TreatmentTPDTPPPSRPDOPChl-aAPABiomasspHDO
AverageField P. crispus area0.140.100.040.040.060.1330.14405.728.208.66
Indoor treatment0.150.110.040.080.030.1324.42101.198.007.24
Field—IndoorF value0.0050.1470.5653.1333.2250.022.01916.0951.7914.429
p value0.9420.7030.4560.0830.0790.8870.16200.1870.041
AverageField Non-aquatic plant area0.130.070.060.050.020.1542.360.008.178.20
Indoor control0.110.050.060.040.030.0715.420.007.365.30
Field—IndoorF value1.6496.6490.0001.71214.5828.876132.922 58.489171.429
p value0.2060.0130.9870.1970.0000.0050.000 0.0000.000
Standard deviationField P. crispus area0.070.050.030.020.040.1112.35258.310.391.80
Indoor treatment0.160.130.040.110.040.1111.1667.390.652.61
Field Non-aquatic plant area0.040.020.020.020.010.087.960.000.330.57
Indoor control0.040.010.040.020.020.042.020.000.270.91
RangeField P. crispus area0.210.150.080.090.120.3140.001000.001.856.37
Indoor treatment0.560.430.120.410.160.2935.00210.652.007.87
Field Non-aquatic plant area0.150.080.100.060.040.2927.000.001.431.98
Indoor control0.170.050.130.070.050.117.000.000.893.68
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Wang, L.; Zhang, L.; Song, H.; Dong, B.; Wang, Y.; Yu, W.; Wu, Y.; Wu, X.; Ge, X. The Effect of the Potamogeton crispus on Phosphorus Changes throughout Growth and Decomposition: A Comparison of Indoor and Outdoor Studies. Sustainability 2023, 15, 3372. https://0-doi-org.brum.beds.ac.uk/10.3390/su15043372

AMA Style

Wang L, Zhang L, Song H, Dong B, Wang Y, Yu W, Wu Y, Wu X, Ge X. The Effect of the Potamogeton crispus on Phosphorus Changes throughout Growth and Decomposition: A Comparison of Indoor and Outdoor Studies. Sustainability. 2023; 15(4):3372. https://0-doi-org.brum.beds.ac.uk/10.3390/su15043372

Chicago/Turabian Style

Wang, Lizhi, Liying Zhang, Hongli Song, Bin Dong, Yun Wang, Wanni Yu, Yuanzhi Wu, Xiaodong Wu, and Xuguang Ge. 2023. "The Effect of the Potamogeton crispus on Phosphorus Changes throughout Growth and Decomposition: A Comparison of Indoor and Outdoor Studies" Sustainability 15, no. 4: 3372. https://0-doi-org.brum.beds.ac.uk/10.3390/su15043372

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