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All issues No 426 (2025) Knowl. Manag. Aquat. Ecosyst., 426 (2025) 13 Full HTML
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Knowl. Manag. Aquat. Ecosyst.
Number 426, 2025
Topical issue on Ecological, evolutionary and environmental implications of floating photovoltaics
Article Number 13
Number of page(s) 14
DOI https://doi.org/10.1051/kmae/2025005
Published online 14 May 2025

© P. Vouhe et al., Published by EDP Sciences 2025

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 ecosystems play a key role in the global carbon cycle as they are the planet’s main reservoir of active carbon (Cole et al., 2007). In fact, in addition to autochthonous organic matter (OM) derived from aquatic biota, they receive large amounts of terrestrial carbon (C) (Drake et al., 2018), commonly referred to as terrestrial subsidies, mainly in the form of organic matter (particulate, dissolved). Once in aquatic ecosystems, this C is transported, stored, transformed, transferred to food webs and/or emitted to the atmosphere as greenhouse gases (GHGs), or returned to terrestrial food webs as aquatic subsidies, such as insect emergence Although research on aquatic and terrestrial ecosystems has long been decoupled, their interactions in terms of mass and energy flows are strong (Soininen et al., 2015). Indeed, terrestrial organic matter plays a central role in the functioning of aquatic ecosystems, influencing both their biodiversity and the ecosystem processes they support. Aquatic biodiversity, supported in part by terrestrial subsidies, influences the functioning and diversity of terrestrial ecosystems (e.g., Bartels et al., 2012). As a result, changes in biodiversity in one ecosystem can propagate across ecosystem boundaries and affect the functioning of other ecosystems (cross-boundary effects). Environmental fluctuations—whether natural (e.g. seasonal variation, heavy rainfall, drought) or anthropogenic (e.g. environmental fragmentation, eutrophication), can alter terrestrial subsidies as well as the dynamics of OM (production, transformation and retention) within aquatic ecosystems. These can significantly alter the structure of aquatic food webs and the fate of organic carbon (OC) in aquatic systems, thereby inducing changes in C fluxes at the interface between aquatic and terrestrial ecosystems, either in the form of GHG or secondary production.

Among anthropogenic sources of disturbance, floating photovoltaic (FPV) power plants consist of photovoltaic panels mounted on floats moored to the surface of a body of water (Sahu et al., 2016). These installations represent a new and rapidly growing use of water bodies, as they produce electricity with 14% higher yields for the same area than land-based power plants (Choi, 2014). FPV plants are generally installed on artificial water bodies (e.g. reservoirs, ponds, gravel pit lakes) and on average cover 34.2% (ranging from 0.004 to 89.9%) of the water surface, with small lakes having proportionally the highest coverage (Nobre et al., 2024). Lakes, especially small ones, not only host high aquatic biodiversity (e.g. Biggs et al., 2017; Heino et al., 2021), but also support terrestrial species such as birds and bats (e.g. Voigt and Kingston, 2016; Torrent et al., 2018; Zamora-Marín et al., 2021). Furthermore, they are a key component of the global C cycle, accounting for 35% of CO2 and 72% of CH4 emissions from lakes worldwide (Downing et al., 2006; Holgerson and Raymond, 2016). Yet, empirical and robust assessments of the ecological impacts of FPV on lakes are still poorly documented (Nobre et al., 2023). FPV implementation may affect OM dynamics, such as the amount of terrestrial OM subsidies and microalgal or phytoplankton primary productivity, and thus the relative importance of allochthonous and autochthonous OM in lake ecosystems. In addition to altering OM dynamics, floating photovoltaics can affect the way OM is processed in aquatic ecosystems. For example, the decreased inputs in terrestrial particulate organic carbon and changes in prevailing conditions (e.g. light, oxygen, temperature) may affect the biomass and community structure of benthic macroinvertebrates, especially leaf-shredding macroinvertebrates and grazers, and thus reduce the rate of OM decomposition and increase OM accumulation in sediments. In addition, reduced dissolved oxygen concentrations and prolonged stratification can promote anoxic conditions that increase CH4 production, particularly when paired with OM accumulation. As a result, FPV may induce changes in the functional links between aquatic ecosystems and their adjacent terrestrial ecosystems, in particular through changes in C fluxes (e.g. terrestrial organic C subsidies to aquatic ecosystems, insect emergence from aquatic to terrestrial ecosystems).

In this review, we synthesise current knowledge on C fluxes at the terrestrial–aquatic ecosystem interface, with the aim of developing a conceptual framework to support the formulation of hypotheses and research questions regarding the impacts of FPV on ecosystem functioning and biodiversity. We focus on C fluxes in the form of terrestrial OM (dissolved and particulate OM), autochthonous OM, emergent insects and greenhouse gases (CO2 and CH4). We address three questions: i) how FPV affect C dynamics within aquatic ecosystems; ii) what are the consequences for aquatic ecosystem biodiversity and functioning; and ii) how aquatic ecosystem responses may propagate to affect terrestrial consumers and climate.

2 Organic matter in aquatic ecosystems

Organic matter (OM) in lakes originates from both terrestrial (allochthonous OM) and aquatic ecosystems (autochthonous OM). Photosynthetic production of OM by algae and aquatic macrophytes constitute the autochthonous OM inputs to lakes. Allochthonous OM originates from terrestrial ecosystems in the watershed (e.g., leaf litter, soil OM, dissolved OM) and is delivered to lakes via lateral transport, groundwater, and inflowing rivers and streams. For any given lake, OM supply can derive to varying degrees from internal primary production vs. terrestrial subsidies such as litterfall and lateral contributions from riparian vegetation. Given the diverse sources, OM can vary in chemical composition and size (e.g., coarse/fine particulate OM, labile/recalcitrant dissolved OM). For example, long C-chain polyunsaturated fatty acids are synthesised and abundant in photosynthetic autotrophs, but are lacking in allochthonous resources (Napolitano, 1999), and they are an important determinant of C-trophic transfer efficiency (Gladyshev et al., 2011; Lau et al., 2014; Mehner et al., 2022). OM in lakes is operationally divided into particulate (POM) and dissolved fractions (DOM). The balance between these forms and their eventual fate is determined by a variety of processes (e.g., adsorption/desorption, aggregation/dissolution) and can be mediated by photochemical processes and biological activity (Derrien et al., 2019). OM inputs into lakes vary in time and space depending on environmental variability, including hydroclimatic conditions, light availability, watershed or water-prevailing conditions (e.g., nutrient loadings, pH, transparency). Accordingly, the distribution and origin of OM differ among and within lake ecosystems. For example, in temperate regions, higher inputs of allochthonous OM are expected in autumn when the leaves fall and after heavy rains, compared to the winter season. In contrast autochthonous production is higher in spring and summer. Within lakes, OM supply from pelagic zones (notably in large lakes) is likely more dependent on phytoplankton production while benthic algae, macrophytes and terrestrial OM should dominate OM pools in littoral areas (Doi, 2009). Benthic regions accumulate OM based on the relative allochthonous, littoral and pelagic contributions. In addition, lake size and depth control the relative importance of allochthonous vs. autochthonous origins of the OM pool, depending on the relative watershed size and surface area-perimeter ratio (Gasith and Hosier, 1976;Wetzel, 1990). In particular, small lakes are more heavily influenced by their watershed and, thus, allochthonous OM supply is more important than in larger lakes (Genkai-Kato and Carpenter, 2005; Leal et al., 2023). In shallow lakes, OM inputs are often dominated by benthic microalgal production or macrophytes (Vadeboncoeur et al., 2008). At finer scales, prevailing conditions such as temperature and solar radiation can vary even within lakes and hence, influence the supply, availability and fate of allochthonous and autochthonous OM. In artificial water bodies such as gravel pit lakes, the age of the lake (the time since plenishment) is positively influencing primary production and water nutrient levels, suggesting that the relative proportion of allochthonous to autochthonous inputs decreases as the lake matures (Colas et al., 2020).

The origin of OM, including the relative importance of allochthonous and autochthonous OM, and its chemical properties can have cascading effects on whole ecosystems, regulates the fraction of OM that settles into sediments, which is subsequently either buried, mineralised, transferred within trophic networks, or emitted to the atmosphere. In particular, OM supply has a critical influence on the structure of the food web and the overall functioning of aquatic ecosystems. OM is a basal source of energy for the various microbial and macrobenthic communities (Taube et al., 2018), determining the potential secondary productivity of lakes. Lake food webs, and ultimately secondary production, have historically been assumed to be predominantly based on autochthonous energy sources (Cummins, 1975), probably because most of the seminal studies in limnology have focused on the pelagic zone of large lakes. Since then, several studies have demonstrated the dominance of terrestrial OM in lake systems in lakes (e.g., (Wilkinson et al., 2013; Guillemette et al., 2017), suggesting that allochthonous OM is an important energy source for lake food webs. In a recent meta-analysis, Leal et al. (2023) evaluated the relative importance of allochthonous versus autochthonous OM sources for freshwater food webs based on stable isotope mixing models from 58 published studies. Contrary to the classical view of the dominance of autochthonous energy sources, they reported a predominance of terrestrial OM in the diet of consumers. Further evidence suggests that allochthonous OM supports lake productivity, including pelagic bacterial production (up to 70% of the production, Kritzberg et al., 2004), zooplankton (Hirama et al., 2022), benthic macroinvertebrates and fishes (Carpenter et al., 2005; Cole et al., 2006), and indirectly through nutrient subsidies to primary producers that increase autochthonous OM production (e.g., Rivera Vasconcelos et al., 2018; Hirama et al., 2022). While the predominance of remains debated and likely varies by lake and context, it is clear that lakes process both OM allochthonous and autochthonous OM. The balance between the two influences the fate of organic carbon, i.e., storage, entry into the food web, release to the atmosphere (e.g. Cole et al., 2007).

3 Fate of organic matter in aquatic ecosystems

The fate of OM in lakes is a critical area of research as it influences C cycling, nutrient dynamics, and overall ecosystem functioning. Lakes serve as an important part of the global C cycle (Cole et al., 2007; Tranvik et al., 2018) because they receive substantial OM inputs from the watershed in addition to within-system production from photosynthesis. Much of this OM settles into bottom sediments, making the benthic zone a hotspot of lake metabolism—particularly in small and/or shallow lakes (Pace and Prairie, 2005). Once on the lake bottom, sediment organic carbon is used by aquatic organisms for biomass synthesis (anabolism) and respiration (catabolism). Decomposition is a major contributor to ecosystem respiration, which is responsible for carbon emissions through GHG (Fig. 1A) and nutrient cycling. It includes both the conversion of OM from large to small particles (e.g., leaf decomposition) and the conversion of complex organic molecules (e.g., carbohydrates) to simpler forms including inorganic constituents (e.g., CO2, N, P) (Findlay, 2021). The process is a complex interplay of physical (e.g., abrasion, physical fractionation), chemical (e.g., leaching, photochemical reaction), and biological processes. Sediment organic matter is processed by a wide range of heterotrophic organisms (prokaryotes and metazoans) whose actions and retroactions and contributions depend on OM composition, size and prevailing conditions. Microbes—such as heterotrophic bacteria, archaea, and fungi—are primarily responsible for the majority of OM decomposition, acting directly and indirectly in the decomposition process. They degrade a wide range of particle sizes (DOM and POM) by producing and secreting extracellular enzymes, but may degrade smaller particles more rapidly owing to increased surface-to-volume ratio and access for cells and extra-cellular enzymes (Moore et al., 2004). In addition, microbial conditioning of OM facilitates fragmentation by benthic macroinvertebrates (e.g., Danger et al., 2012; Colas et al., 2016). Benthic macroinvertebrates (e.g., aquatic insects, crustaceans) also play an important role in the fate of sedimentary OM and ultimately in the biogeochemical cycle at the sediment–water interface. These organisms use organic matter and associated microbial biomass as a food resource and release small particles (e.g., FPOM, amino acids) that can be easily incorporated by microbes. Additionally, benthic macroinvertebrates can influence the physical structure and chemical properties of the sediment-water interface. In particular, through bioturbation and bioirrigation, they promote OM availability and oxygen diffusion into the sediments (e.g., Fiskal et al., 2021) modifying benthic bacterial communities, regulating OM decomposition and ultimately GHG production (e.g., Mermillod-Blondin, 2011; Chakraborty et al., 2022). For example, methane (CH4) oxidation in surface sediments is often stimulated by O2 input from chironomid larvae, which enriches populations of methane-oxidising bacteria in the larval tubes and surrounding sediment (“microbial gardening”) (Kajan and Frenzel, 1999). Microbes and detritivorous macroinvertebrates use OM as a source of carbon and energy for their own metabolism and are in turn consumed (e.g., by protozoa for microbes, by fish for detritivores), contributing to the transfer of organic carbon along the “brown food chain”. In addition, the inorganic compounds released during the mineralisation of OM (i.e., nitrogen, phosphorus, carbon dioxide) support primary producers (e.g., phytoplankton, benthic algae), thus enhancing production at higher trophic levels through the “green food chain” (Hirama et al., 2022).

The rate and pathway at which biological communities process organic matter is controlled by its inherent properties (chemical, physical) and the environmental conditions (e.g., dissolved oxygen, pH, temperature and nutrients). In particular, dissolved oxygen (DO) concentration is a key determinant of the microbial community structure and metabolic pathways involved in OM mineralization (e.g., Bastviken et al., 2004; Liu et al., 2021). Under oxic conditions, aerobic respiration predominates, resulting in the release of carbon dioxide (CO2) and inorganic nutrients, until DO is depleted. Subsequently, nitrate reduction (nitrification), manganese reduction, iron reduction, and sulphate reduction—take place, which also generate CO2. In the absence of DO and other alternative inorganic electron acceptors, such as nitrate (NO3−), ferric iron (Fe3+), and sulphate (SO42−), OM decomposition produces CH4, a highly potent greenhouse gas, through the process of methanogenesis, which accounts for 10–50% of the total carbon mineralisation in lakes (Bastviken et al., 2008). Methane is produced by the methanogenic archaea (methanogens, Rudd and Hamilton, 1978), a ubiquitous group in anaerobic environments (anoxic water column, sediments) that decompose organic matter by fermentative microbes (e.g., CO2, hydrogen, acetate, methylated compounds). The final respiration products, including GHGs (CO2, CH4), may dilute and diffuse in the water column or diffuse into the top layer of the sediment (Fig. 1A). During this process, CH4 can be reoxidised to CO2 by methanotrophs under oxic and anoxic conditions (see review of Guerrero-Cruz et al., 2021). Methanotrophs can oxidise up to 99% of the CH4 produced in freshwater lakes (Bastviken et al., 2008), thereby reducing CH4 emissions. In turn, methanotrophs can be an important carbon source for sedimentary macroinvertebrates (e.g., chironomid larvae), especially in eutrophic lakes (Fiskal et al., 2021), and for zooplankton in the water column (Kankaala et al., 2007; Zha et al., 2021). GHGs that escape oxidation (CH4) and uptake by primary production (CO2) are emitted to the atmosphere through molecular diffusion at the water-air interface, assuming that their partial pressure in water exceeds that in the atmosphere (Fig. 1A). Once produced, CH4 can also be transported from the anoxic sediment layers to the atmosphere by bubble fluxes released from the sediments and by plant-mediated transport of vascular plants (Bastviken et al., 2023). Lakes are estimated to emit approximately 0.6 Pg carbon per year to the atmosphere in the forms of CH4 and CO2 (Tranvik et al., 2009; Holgerson and Raymond, 2016), making them a significant component of the global carbon (C) cycle. Due to their morphology, productivity and connectedness with the watershed, small lakes contribute to 35 % of the CO2 and 72 % of the CH4 emissions from lakes worldwide (Holgerson and Raymond, 2016).

thumbnail Fig. 1

Schematic overview of the GHG dynamic and C fluxes from aquatic ecosystems according to lake type (A) and potential short-term impact of FPV (B). A. Once in the sediment, the processing of organic matter produced CO2 and CH4 (1). Once produced, the oxidation of CH4 to CO2 by methanotrophs and the uptake of CO2 by primary producers mitigate GHG emissions to the atmosphere (2,3). GHGs that escape oxidation and the carbon sink are emitted to the atmosphere by diffusing across the water-atmosphere interface (4). At shallow depth, CH4 is transported from the sediment to the atmosphere through ebullition (5). During the ebullition process, there may be dissolution and oxidation of CH4 into the water column (6). Vascular plants such as macrophytes act as important conduits for CH4 to the atmosphere (7). Emergent aquatic insects, which spend their larval stages in lake sediments and emerge as adults, represent C fluxes from aquatic to terrestrial habitats, providing high quality resources for terrestrial consumers (8). B. Irrespective of lake type, the increased OM supply resulting from the death of primary producers in response to severe shading and from allochthonous inputs could fuel benthic respiration processes and thus GHG production. In particular, CH4 production would increase, favoured by anoxic conditions in the sediments. At shallow depths, much of the CH4 produced would be emitted to the atmosphere via ebullitive pathways. The GHG not emitted by ebullition would accumulate in the water column due to reduced CH4 oxidation and CO2 uptake by primary producers combined with reduced gas transfer velocities beneath the panels. For stratified lakes, high emissions could occur during mixing events, although though FPV would shorten stratification and reduce mixing depths. FPV would reduce insect larval biomass and ultimately adult emergence due to reduced DO, cooler temperature and loss of habitat complexity and trophic resources. Such effects on C fluxes would increase in magnitude with panel coverage, especially in shallow and small lakes, due to higher primary and secondary production prior to FPV implementation.

4 Abiotic controls of organic matter in lakes

The fate of organic matter in lakes is controlled by several factors, including OM inputs (composition and quantity), temperature, oxygen availability, nutrient concentrations and hydrodynamics. The composition and quantity of OM are key factors controlling the fate of OM in lakes and hence the C cycle. The metabolic fate of metabolic processing of OM can shift depending on the composition of OM (Guillemette and Del Giorgio, 2012). In particular, the degree of recalcitrance influences the composition of the microbial community and the metabolic pathways of OM decomposition. Autochthonous OM, which is more labile, is typically decomposed by fast-growing microbial taxa such as bacteria and fungi, and decays faster than allochthonous OM, leading to a high rate of microbial respiration and GHG production (e.g., Bartosiewicz et al., 2021). In contrast, the degradation of allochthonous OM requires the support of microbial communities that specialise in the break-down of more complex, lignin and that exhibit slower growth rates, resulting in slower overall decomposition rates (e.g., Grasset et al., 2021). Moran and Hodson (1990) estimate that only 1–10% of allochthonous OM in sediments is readily utilised and degraded by bacteria. As a result, recalcitrant OM tends to be buried and dominates lake sediments (e.g, West et al., 2012; Guillemette et al., 2017; Grasset et al., 2018), where it either accumulates... or is gradually transformed into more labile organic compounds that can be used as a C sources by microbes. Both organic matter sources support GHG production, but exhibit different decomposition dynamics (i.e., rapid decomposition for autochthonous material and progressive decomposition for allochthonous material) (e.g., West et al., 2016; Grasset et al., 2018). Ultimately, the quantity of organic matter in the form of OM, whether as DOC or POM, entering lakes significantly influences the carbon cycle by increasing both CO2 (e.g., Karlsson, 2007) and CH4 production (e.g., (Gudasz et al., 2015; Berberich et al., 2020). In particular, Grasset et al., (2021) provide evidence that CH4 production increases linearly with POM deposition, regardless of the composition of organic matter.

Several factors influence directly and indirectly the availability of allochthonous and autochthonous OM in lakes, including precipitation, solar radiation, nutrients concentrations and temperature. In particular, solar radiation and high nutrient levels enhance primary production, increasing autochthonous OM input to the sediments, as well as hypoxic or anoxic conditions that promote anaerobic decomposition pathways. Temperature affects organic carbon substrates, but also the composition of the microbial communities and microbial enzyme activities (e.g., Yang et al., 2023). Although the decomposition of allochthonous OM is likely more temperature-sensitive than that of easily decomposed OM (Bergström et al., 2010), higher temperatures increase microbial activity, leading to faster aerobic and anaerobic decomposition rates of OM into CO2 and CH4 (e.g., Yvon-Durocher et al., 2010; Gudasz et al., 2015). In their study, Praetzel et al. (2020) reported that with a temperature increase of 10°C, production rates of CO2 doubled, and those of CH4 were even by a factor of 2 to 11. Temperature also influences the composition of microbial communities in sediments. For example, warm temperature increases the abundance of methanogenic archaea, leading to increased CH4 production (Solomon et al., 2015). In addition, temperature affects the physical structure of lakes, particularly through its influence on thermal stratification, which creates distinct thermal layers, including a warm, oxygen-rich epilimnion and a hypolimnion. Higher temperatures can strengthen thermal stratification, reducing oxygen availability in deep waters and increasing the relative importance of anaerobic processes and the potential for organic matter preservation in lake sediments (Jenny et al., 2016). Lake mixing, driven by temperature and wind changes, influences the distribution and transport of OM. For example, in well-mixed lakes, OM is distributed more evenly throughout the water column, while in stratified lakes, OM may accumulate in specific layers, such as the hypolimnion or sediments. Periodic mixing events, such as turnover in temperate lakes, can reintroduce OM and nutrients from the sediments to the surface waters, affecting lake productivity and GHG production (e.g., Vesala et al., 2006; Åberg et al., 2010). This effect is also common in shallow lakes that exhibit temporary thermal stratification (Martinsen et al., 2019; Davidson et al., 2024).

5 Carbon effluxes mediated by animal movements in lakes

Ecosystems are rarely isolated from each other. As described above, aquatic ecosystems receive OM subsidies from adjacent terrestrial ecosystems which they dynamically process, transport, or sequester, contributing significantly to the global carbon budget and influencing climate. Supported by the allochthonous OM, aquatic ecosystems harbour high biodiversity, which can influence food webs and ecosystem functioning beyond their boundaries. Although mass and energy fluxes between terrestrial and aquatic ecosystems are quantitatively skewed towards the terrestrial-to-aquatic direction, aquatic-to-terrestrial exports are of higher biological quality (Bartels et al., 2012; Twining et al., 2021). For example, studies show that emergent insects are an important source of polyunsaturated fatty acids (Twining et al., 2021; Fehlinger et al., 2023) essential to the nutritional requirements, development, and reproductive success. These aquatic subsidies are biologically mediated and depend on the movement of animals (Schindler and Smits, 2017). Two types of movement are involved: emergence of aquatic organisms into terrestrial ecosystems during adult life stages (mostly insects, but also amphibians), and aquatic organisms that are preyed upon by terrestrial predators and removed from aquatic ecosystems. In particular, emerging adult aquatic insects and amphibians constitute important subsidies for a wide range of terrestrial consumers, including birds (Nakano and Murakami, 2001) and bats (Larsen et al., 2016) (Fig. 1A). Global attempts to quantify such fluxes are scarce in the scientific literature. Nevertheless, an increasing number of articles have been published in recent years, particularly on river subsidies to surrounding terrestrial ecosystems, pointing to the importance of aquatic-terrestrial fluxes mediated by emergent insects (Schindler and Smits, 2017). Lake-land linkages are less well studied, but there is ample evidence that the amount of insects produced can be substantial, with some lakes in Iceland producing up to 1,000 tonnes of midges dry mass per year (Hoekman et al., 2011), comparable to river contributions. For example, Bartrons et al. (2013) summarised data from 28 temperate lakes and 36 stream ecosystems, and insect export production ranging from 0.2 to 34.6 g C m−2 yr−1 in lakes and from 0.3 to 34.6 g C m−2 yr−1 in streams. Notwithstanding, Gratton and Zanden (2009) developed a conceptual model to assess the transfer of aquatic production to adjacent terrestrial ecosystems via emerging insects and suggested that standing waters may export more carbon than from flowing waters due to a more favourable ratio of aquatic habitat area to perimeter in lakes. Based on a further development of this model, Bartrons et al. (2013) estimated that up to 79% of total emerging insect production in Wisconsin (USA) depends on lentic ecosystems.

This substantial emergence has direct impacts on terrestrial decomposer communities and terrestrial predators, enhancing growth rates, reproductive success, and survival, and with effects that can persist for up to two years. For decomposers, the addition of aquatic midges has been shown to increase terrestrial arthropod biomass by 68% after one year and by 108% after two years (Hoekman et al., 2011). For a wide range of terrestrial predators, represents an important food source. They can contribute more than half of the diet of insectivorous bats and birds (Recal de et al., 2021), and may constitute their entire diet in late spring (Piovia-Scott et al., 2016). Such resources are particularly important during the breeding season (Kunz and Fenton, 2003). For example, studies have shown that bird species that nest near water time their reproductive efforts to coincide with peaks in insect emergence, ensuring abundant prey for their offspring (Baxter et al., 2005). Some studies have reported that female bats make more visits to water during lactation (Adams and Hayes, 2008) and that female nursery roosts tend to be closer to water than male roosts (Encarnação et al., 2005). In addition to serving as food, emergent insects contribute to nutrient and carbon cycling in riparian and adjacent terrestrial zones after they die and decompose in soils (Gratton and Zanden, 2009). Aquatic subsidies by emergent insects is influenced by both environmental and biotic factors within aquatic ecosystems and can also be mediated by terrestrial predators. For example, while fish exert top-down control on aquatic food webs (e.g. Matveev and Robson, 2014), thereby controlling the biomass of emergent insects, they are themselves preyed upon by a wide variety of terrestrial vertebrates such as reptiles, mammals and birds. Other factors that influence community composition and the abundance of benthic larval forms may ultimately influence adult emergence patterns, including microhabitat availability (Merten et al., 2014), particularly the presence of coarse organic matter and submerged aquatic vegetation (e.g., Schad et al., 2020), water level (e.g., Drummond et al., 2015), dissolved oxygen (DO) concentration (e.g., Connolly et al., 2004; Martin-Creuzburg et al., 2017), water temperature (e.g., Watanabe et al., 1999), and water quality (e.g., Schmidt et al., 2013). As these environmental factors tend to vary with lakes bathymetry, Several studies have reported the highest insect emergence rates in shallow areas, such as littoral zones, compared to deeper depths (e.g. Martin-Creuzburg et al., 2017; Mathieu‐Resuge et al., 2021).

6 Potential cross-ecosystems effects of floating photovoltaics on carbon fluxes

- Effects of FPV mediated by changes in OM dynamics

FPV can affect C fluxes across aquatic–terrestrial boundaries through changes in the availability and distribution of organic matter, which can lead to cross-ecosystem effects on biodiversity and ecosystem functioning (Fig. 1B). The short-term effect of FPV implementation would be an increase in both terrestrial subsidies to aquatic ecosystems and autochthonous organic matter settling onto the lake bottom. In particular, shading of the water surface by FPV can lead to large declines in phytoplankton and macrophyte populations (e.g., Köhler et al., 2010; Bergström and Karlsson, 2019), increasing the supply of authochthonous OM to bottom sediments (Fig. 2A). This initial supply of autochthonous OM to the bottom sediments is expected to be greater with increasing panel coverage and for shallow lakes where the lake bottom receives sufficient sunlight to stimulate aquatic macrophyte growth. Implementation works (e.g., riparian cutting, shredding vegetation, landscaping) would also increase the supply of allochthonous OM to aquatic ecosystems (Fig. 2B), mainly from leaching of fine and dissolved organic carbon from the surrounding landscape (Giling et al., 2015). Such inputs can be significant, especially when terrestrial photovoltaic installations are located adjacent to FPV systems. Long-term effects on OM availability and distribution are difficult to predict, as they depend on factors such as the extent of riparian disturbance (e.g., extent of deforestation, replanting of vegetation, panel coverage), nutrient loads or community dynamics (e.g., hysteresis). As a result, different scenarios are possible for both allochthonous and autochthonous OM (Figs. 2A and B), which can occur simultaneously, resulting in different scenarios of balance between the two OM sources, depending on the initial trophic state of the lake. In particular, primary producer biomass including phytoplankton and macrophytes can remain consistently lower over time than before implementation (Fig. 2A, S1auto) due to shading effects combined with changes in water temperature (cooler temperature reducing metabolic rates) and lake hydrodynamic (i.e. reduced continuous and cumulative stratification duration) (Château et al., 2019; Haas et al., 2020; Exley et al., 2021). However, while severe shading could inhibit the growth of submerged macrophytes (e.g., Gao et al., 2021), some macrophytes may exhibit phenotypic responses to adapt to low light. For example, macrophyte species with erect stems may compensate for light limitation by increasing internodal growth towards the water surface not covered by leaves (Middelboe and Markager, 1997). On the other hand, rosette growth forms are more sensitive to light limitation, so that species with this growth form are more likely to disappear after FPV installation (Vestergaard and Sand-Jensen, 2000). In addition, low light can also lead to changes in the composition and dominance of primary producer communities, which can result in increased phytoplankton densities due to the released competition with in shallow lakes, including low-light-adapted taxa such as cyanobacteria (i.e., Schwaderer et al., 2011; Yamamichi et al., 2018) (Fig. 2A, S2auto).

This paradox could result in different primary producer densities over time that may be lower (S2b auto), equivalent (S2a auto) or higher (S2c auto) than densities before implementation. For example, FPV may create a new habitat for the development of primary producers adapted to low light conditions (Nobre et al., 2023), as reported by de Lima et al. (2021) nine months after implementation, supporting the latter scenario. Riparian vegetation acts as a filter for leaching materials from the surrounding landscape (Lowrance, 1998). As a result, cutting of riparian vegetation may increase the supply of fine and dissolved organic carbon (Fig. 1B, S1allo) to aquatic ecosystems. Ultimately, however, cutting of riparian vegetation would result in a decrease in terrestrial subsidies (Webster et al., 1990; France et al., 1996), either in the form of coarse organic matter or dissolved organic matter (Fig. 1B, S2allo). For example, France et al. (1996) calculated a 40-fold reduction in terrestrial DOC subsidies to oligotrophic lakes following riparian removal. Yet, the presence of a buffer strip may offset the decrease in terrestrial OM to aquatic ecosystems (Kiffney and Richardson, 2010). Over time, terrestrial inputs are expected to increase as vegetation regrows (S2b allo).These changes in autochthonous and allochthonous OM supply would result in different balances between the two sources and hence different ecosystem metabolisms, including regime shifts compared to pre-FPV ecosystem metabolism. For example, in eutrophic ecosystems, ecosystem metabolism is mainly driven by autochthonous OM relative to allochthonous OM, resulting in net autotrophy. In such an ecosystem, reduced biomass of primary producers due to light limitation would result in a switch towards net heterotrophy if coupled with a higher supply of allochthonous OM than endogenous production. Conversely, in oligotrophic lakes, allochthonous OM regularly exceeds primary production, leading to net heterotrophy. But some studies have found that even oligotrophic ecosystems can be heterotrophic, they can be also autotrophic or a combination, depending on DOC concentration (Prairie et al., 2002; Hanson et al., 2011), making predictions on the ecosystem trends after the installation of FPV highly uncertain. For example, DOC concentrations may be of terrestrial or autochthonous origin. Furthermore DOC degradation can be driven by sunlight (i.e. photochemical processes), especially in surface waters. Low light due to FPV implementation may reduce DOC photodegradation, resulting in increased concentrations in the water column that contribute to the decrease in photosynthetic radiation and thus a reduced primary productivity (e.g., Brett et al., 2017). Considering a prevalent heterotrophic regime in oligotrophic lakes, we can hypothesize a decrease in coupled with increased primary producer biomass, could enhance lake productivity and a regime switch towards autotrophy. Regime shifts may occur rapidly after FPV implementation and then stabilise over time due to maintenance of shading conditions or the presence of hysteresis, or they may switch again depending on changes in prevailing conditions (OM supply and reactivity, nutrients loads, community dynamics).

- Effects of FPV on GHG production and emissions

The changes in the OM pool and lake metabolism resulting from the implementation of FPV could ultimately disrupt carbon (C) dynamics and increase GHG emissions from lakes. However, the magnitude of this impact may vary based on: 1) the bathymetry and hydrodynamics of the lake, which influence the fate of OM (Figs 1B, and 2) the specifics of the FPV implementation (such as the panel coverage and the technology used), which modulate the extent of impact, with strong interactions among them. For example, FPV-induced changes in temperature and oxygen could ultimately alter the mixing and stratification regime (e.g., Exley et al., 2022) and thus, the fate of OM in lakes. Furthermore, the magnitude of changes in OM dynamics is likely to be greater in small lakes, as FPV coverage tends to be greater in proportion to surface area (Nobre et al., 2024) and, as small lakes are typically more productive. Two thirds of the world’s FPVs are installed in small water bodies (Nobre et al., 2024), such as gravel pit lakes and ponds. However, some FPVs already impact large and stratified water bodies such as reservoirs (e.g. the Alqueva FPV in Portugal). This is likely to increase in the future, as suggested by recent calls to implement FPVs on hydropower reservoirs as a means to reduce the global energy C footprint (Almeida et al., 2022; Woolway et al., 2024). Consequently, the effects of FPV on GHG production and emissions would depend on lake size, hydrodynamics, and stratification, with more pronounced effects in smaller, shallow lakes where FPV coverage represents a larger proportion of the surface area. In contrast, larger lakes, with proportionally lower coverage of the surface water, are likely to experience weaker effects. Irrespective of the type of water body, once in the bottom sediment of FPV lakes, OM may exhibit a potentially antagonistic fate. On one hand, a high supply of organic matter, especially labile autochthonous OM derived from macrophytes mortality in shallow depths to the bottom sediments can stimulate microbial respiration, especially methanogenesis, as FPV reduces DO availability (e.g., Armstrong et al., 2020; Yang et al., 2022; Ray et al., 2024). Several studies have reported a strong correlation between lake productivity and methanogenesis rates (e.g., West et al., 2016; Colas et al., 2020), as anoxic conditions coupled with high OM supply in productive lakes promote methanogenesis and thus methane emissions to the atmosphere, especially in shallow lakes (Colas et al., 2020). On the other hand, lower temperatures below the FPVs (e.g. Ji et al., 2022; Nobre et al., 2025) may slow down metabolic processes that scale with temperature, i.e. methanogenesis and mineralisation of organic carbon, reducing GHG production and promoting C burial. Only one study examined the effect of FPV on GHG production and emissions in shallow ponds (Ray et al., 2024) and reported twice the average GHG production related to macrophyte decomposition under reduced oxygen conditions (despite lower temperatures). In deep lakes, we can probably assume that although colder temperatures will slow down metabolic activities, the addition of OM combined with DO reduction will ultimately lead to higher GHG and especially, CH4 production, as already demonstrated for deep lakes undergoing eutrophication (e.g., DelSontro et al., 2018). However, such an increase in GHG production is expected to be greater for shallow lakes due to a higher OM pool derived from primary producers, favored by the lack of light limitation on the bottom sediment prior to FPV coverage. In addition, the input of autochthonous and allochthonous OM resulting from the short-term effects of FPV implementation may stimulate the mineralisation of pre-existing sedimentary carbon in lake sediments through positive priming effects (e.g. Yang et al., 2023), which could enhance respiration processes and ultimately GHG production.

Once GHGs are produced, the reduced DO concentrations and primary producers could dampen CH4 oxidation by methanotrophs and photosynthesis. Both processes strongly influence the proportion of GHGs that ultimately reach the atmosphere, especially in deep lakes. In addition, FPV may profoundly reduce water–atmosphere gas exchange and turbulent diffusion by reducing stratification stability, mixing depths of deep lakes (e.g., Exley et al., 2022; Ji et al., 2022; Liu et al., 2023) and near-surface wind speeds (e.g. Ilgen et al., 2023; Liu et al., 2023) and hence gas transfer velocities (Ray et al., 2024). This may enhance water column stability and, when combined with increased respiration and reduced photosynthesis and methanotrophy, result in a greater proportion of the GHGs produced in the sediments being stored in the water beneath the panels. Consequently, GHG emissions may not increase in direct proportion with GHG production after FPV implementation, although episodic release events may still occur during disturbances that affect lake hydrodynamics (e.g. lake mixing) and water-atmosphere exchange (e.g. rain, crosswinds, FPV operation). However, the highest methane production in the anoxic sediment layers could ultimately lead to increased transport of CH4 to the atmosphere by ebullition, especially in shallow waters (DelSontro et al., 2016; Ray et al., 2024).

Predicting the long-term effects of FPV remains difficult due to the lack of empirical data at high spatial and temporal resolution and the heterogeneity of water bodies (e.g. lake size and depth, trophic status) and technologies used (e.g. panel coverage). Yet, it is likely that GHG dynamics and emissions will vary concomitantly with the whole-lake metabolism, which is shaped by the balance between allochthonous and autochthonous OM, along with DO and temperature dynamics. In particular, a metabolic shift from autotrophy to heterotrophy may reduce CO₂ emissions but increase CH4 emissions, as already described for shallow gravel pit lakes (Colas et al., 2020). Conversely, a switch from autotrophic to heterotrophic lakes would reduce C emissions in the form of CH4 but increase CO2 emissions. In addition, such a switch may enhance C storage in sediments, especially under conditions of allochthonous OM dominance and lower temperatures, which reduce metabolic activities associated with OM catabolism.

- Effect of FPV on C flux from insect emergence

Across different lake types, FPVs may significantly reduce carbon (C) fluxes via emergent insects by affecting both the distribution of benthic larval forms and the emergence of adult insects (Fig. 1B). The availability of OM, particularly submerged aquatic vegetation, algae, and coarse detritus (e.g., leaf litter), plays a crucial role in shaping trophic resources and habitat for insect larvae. Consequently, any alterations in these resources and habitats could impact adult emergence patterns. Macrophytes, for example, increase habitat complexity by providing shelter and food sources for various insect larvae (Phillips et al., 2016). As a result, lakes with dense and diverse vegetation generally support a more diverse community of aquatic insects, including Trichoptera and Diptera (e.g. Heino, 2008; Thomaz and Cunha, 2010). Likewise, the availability of coarse detritus (e.g. leaf material) plays an important role in supporting the diversity of macroinvertebrates by providing both habitat and food resources (e.g. Kovalenko et al., 2012; Wissinger et al., 2021). In oligotrophic lakes, external inputs of coarse detritus may provide a critical nutrient source that sustains detritivore-based food webs, a pattern well-documented in streams (Richardson, 1992). Environmental conditions, particularly dissolved oxygen concentrations, temperature, and water quality strongly influence a key role in determining the distribution of insect larvae and adult emergences. Low DO levels impair growth and ultimately reduce insect abundance. For example, using a mesocosm approach, Connolly et al. (2004) reported that hypoxia significantly reduced the survival of several macroinvertebrate species such as mayflies and caddisflies. Moreover, decreased oxygen saturation led to suppressed emergence across multiple insect taxa, increasing mortality and inducing sublethal stress that hindered insect development. Water temperature is also an important determinant of aquatic insects, affecting their survival, physiology, life cycles, activities and population dynamics. Most studies have focused on the effects of increased temperature because of the urgent need to predict the effects of global warming on aquatic biodiversity. Accordingly, an increase in insect growth and development was observed with increasing temperature up to a certain threshold, beyond which performance declines due to thermal stress (e.g., Ward and Stanford, 1982; Sweeney et al., 1992). reduced temperatures following FPV installation in temperate lakes implementation (Nobre et al., 2025) is expected to slow insect metabolism and delay their emergence compared to lakes without FPV, resulting in less synchronized emergence events (e.g., Harper and Peckarsky, 2006; Hassall and Thompson, 2008; Bonacina et al., 2023). Once emerged, adults could be attracted to FPVs by polarisation, resulting in reproductive failure of eggs laid on artificial surfaces and death from exhaustion (Horváth et al., 2010; Black and Robertson, 2020). These potential changes in emergent insect biomass and phenology may subsequently affect terrestrial predators, such as bats and birds, by reducing prey availability and decoupling predator–prey interactions (i.e. phenological mismatch), thereby disrupting food webs (Hansson et al., 2014). In addition to reduced prey availability, a decline in insect emergence could diminish the export of essential fatty acids and carbon from aquatic to terrestrial ecosystems, particularly in small and/or shallow lakes ((Martin-Creuzburg et al., 2017). As a result, organisms that rely on these nutrient-rich subsidies, including terrestrial invertebrates, amphibians, and insectivorous birds, may experience nutritional deficiencies, potentially altering their reproduction, growth, and survival. These cascading effects could further weaken the ecological linkages between aquatic and terrestrial food webs, ultimately disrupting ecosystem functioning at multiple trophic levels. In addition to limiting aquatic subsidies FPVs can also create a ’lake effect’, increasing the risk for bird of collisions with their structures (Benjamins et al., 2024). Additionally, FPVs also alter waterbird habitats, leading to a reduction in diversity and changes in assemblages. Observed impacts include a decrease in the proportion of diving birds (Song et al., 2024) and altering the top-down control of terrestrial predators on fish communities ultimately affecting aquatic food webs. Similar to their impact on GHG emissions, FPVs influence on insect emergence would also vary with lake morphometry (Fig. 1B) and panel coverage. Research indicates that insect biomass decreases with increasing lake depth (Martin-Creuzburg et al., 2017; Mathieu‐Resuge et al., 2021), likely due to lower DO levels and reduced submerged vegetation. Consequently, littoral areas and shallow lakes support higher insect abundance than pelagic areas and deeper lakes. As a result, the effects of FPV on insect emergence may be stronger in shallow lakes and intensify as panel coverage increases. Such effects of FPV on insects emergences and hence aquatic subsidies to terrestrial consumers could persist alongside their shading effects on autochthonous OM, lower water temperature and reduced DO levels. Even though FPV do not typically cover littoral areas, environmental changes they induce, especially at high coverage, could still impact these habitats and, in turn, insect biomass and emergence. For instance, FPV may cause changes in water quality by releasing nutrients and contaminants from anoxic sediments (e.g., Yang et al., 2022; Liu et al., 2023). When combined with the lack of light limitation and the sheltering effect of FPVs, these conditions could promote blooms of primary producers, leading to more eutrophic littoral areas. Such changes may negatively affect insect populations and contribute to increased CH4 production. In addition, corrosion and degradation of FPVs over time could release microplastics (e.g. Pouran et al., 2022), which may have negative effects on the growth, reproduction and survival of insect larvae, ultimately shortening emergence times and decreasing the number of emerging insects (Ribeiro-Brasil et al., 2022).

thumbnail Fig. 2

Conceptual model predicting short- and long-term effects of FPV implementation on sediment organic matter inputs from autochthonous production (A) and terrestrial sources (B). Red arrow represents time of FPV implementation.

7 Future directions and applications

In this review, we synthesised current knowledge on carbon (C) fluxes at the interface between aquatic and terrestrial ecosystems, with a particular focus on lakes. We highlighted the role of OM dynamics, including terrestrial OM, in supporting lake productivity and functioning, and the cascading effects on climate through greenhouse gas emissions and on terrestrial biodiversity via aquatic subsidies. By altering OM dynamics − such as terrestrial subsidies, autochthonous production, and OM catabolism and anabolism − FPV have the potential to trigger cross-ecosystem impacts. These changes in C fluxes would vary in magnitude and strength depending on factors like initial trophic state, lake morphometry and panel coverage. We primarily considered terrestrial OM as the main allochthonous source. However, FPV systems, constructed from plastic materials, may introduce various plastic contaminants into aquatic environments. The long-term durability of plastic floaters in water remains uncertain. In addition to their potential detrimental effects on insect larvae and emergence, may act as a novel allochthonous OM source entering aquatic ecosystems, contributing to the OC pool in sediments and, consequently, influencing GHG production and emissions (Zhang et al., 2022). FPV technology is often presented as a sustainable solution for meeting energy demands, particularly by reducing the carbon footprint of electricity generation and mitigating the effects of global warming on lakes (e.g. Almeida et al., 2022; Exley et al., 2022; Woolway et al., 2024). We argue that such claims should be evaluated cautiously, in the absence of long-term studies using robust before–after control–impact designs (Nobre et al., 2024). Without a full quantification of C dynamics, including C emissions and sequestration, accurately predicting the carbon footprint of FPV systems remains challenging. Beyond their potential impact on climate through greenhouse gas emissions, FPV installations could also significantly affect terrestrial consumers such asbats and birds, as well as the productivity of nearby terrestrial ecosystems by reducing aquatic subsidies aquatic subsidies such as insect emergence. Research on these cross-ecosystem effects is still lacking. Future studies should investigate how a decline in insect emergence following FPV implementation may lead to losses in consumer diversity and activity. lake morphometry, trophic state, regional climate, and FPV coverage are likely to interact in mediating the effects of FPVs on C dynamics and fluxes, making ecological impacts difficult to predict. This complexity underscores the need for monitoring programmes that identify how these factors interact and mediate the effects of FPV on C dynamics and fluxes, to provide guidance to reduce potential negative impacts of FPV on aquatic and terrestrial productivity and the global C cycle. The impact of FPV on C dynamics and fluxes at the interface between aquatic and terrestrial ecosystems is expected to be greater in small and shallow water bodies. These ecosystems typically exhibit high primary and secondary productivity, strong connections with adjacent terrestrial ecosystems, and a high potential for FPV implementation with high panel coverage (Nobre et al., 2024).

Acknowledgements

The authors thank all scientists working to advance to provide empirical knowledge on the ecological effects of power plants, and in particular the scientists involved in the SOLAKE, FLOATIX and SOLFLUX research projects. We would like to thank the organisations funding the SOLAKE and FLOATIX projects (OFB, ADEME) and the SOLFLUX research project (OFB, CNR). Finally, we would like to thank the journal editors for inviting this submission, and the two anonymous reviewers whose constructive suggestions improved the quality of the manuscript.

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Cite this article as: Vouhe P, Rodríguez-Pérez H, Wissel B, Lengagne T, Metaireau A, Colas F. 2025. Potential impacts of floating photovoltaics on carbon fluxes across aquatic-terrestrial boundaries. Knowl. Manag. Aquat. Ecosyst., 426, 13. https://doi.org/10.1051/kmae/2025005.

All Figures

thumbnail Fig. 1

Schematic overview of the GHG dynamic and C fluxes from aquatic ecosystems according to lake type (A) and potential short-term impact of FPV (B). A. Once in the sediment, the processing of organic matter produced CO2 and CH4 (1). Once produced, the oxidation of CH4 to CO2 by methanotrophs and the uptake of CO2 by primary producers mitigate GHG emissions to the atmosphere (2,3). GHGs that escape oxidation and the carbon sink are emitted to the atmosphere by diffusing across the water-atmosphere interface (4). At shallow depth, CH4 is transported from the sediment to the atmosphere through ebullition (5). During the ebullition process, there may be dissolution and oxidation of CH4 into the water column (6). Vascular plants such as macrophytes act as important conduits for CH4 to the atmosphere (7). Emergent aquatic insects, which spend their larval stages in lake sediments and emerge as adults, represent C fluxes from aquatic to terrestrial habitats, providing high quality resources for terrestrial consumers (8). B. Irrespective of lake type, the increased OM supply resulting from the death of primary producers in response to severe shading and from allochthonous inputs could fuel benthic respiration processes and thus GHG production. In particular, CH4 production would increase, favoured by anoxic conditions in the sediments. At shallow depths, much of the CH4 produced would be emitted to the atmosphere via ebullitive pathways. The GHG not emitted by ebullition would accumulate in the water column due to reduced CH4 oxidation and CO2 uptake by primary producers combined with reduced gas transfer velocities beneath the panels. For stratified lakes, high emissions could occur during mixing events, although though FPV would shorten stratification and reduce mixing depths. FPV would reduce insect larval biomass and ultimately adult emergence due to reduced DO, cooler temperature and loss of habitat complexity and trophic resources. Such effects on C fluxes would increase in magnitude with panel coverage, especially in shallow and small lakes, due to higher primary and secondary production prior to FPV implementation.
In the text
thumbnail Fig. 2

Conceptual model predicting short- and long-term effects of FPV implementation on sediment organic matter inputs from autochthonous production (A) and terrestrial sources (B). Red arrow represents time of FPV implementation.
In the text

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