Heat Recovery as a Tool for Reducing the Thermal Impact of Effluents from Wastewater Treatment Plants

Abstract

Water temperature is a key ecological and metabolic factor in rivers and other continental systems, and thermal pollution caused by anthropogenic activities (dams, discharges, urban stormwater, industrial cooling) alters the natural thermal regime of rivers, modifying the structure and functioning of communities (primary producers, macroinvertebrates and fish) and favouring thermophilic and often invasive species. Wastewater treatment plants (WWTPs) generate and discharge excess heat: their effluents are often several degrees above the temperature of the receiving river, which increases the metabolism of communities, favours eutrophication and can intensify the effects of nutrients and toxic pollutants. This excess heat from wastewater is a major renewable energy resource that can be recovered using heat pumps, both in buildings and in the treatment plants themselves, as well as in district heating networks, reducing the demand for fossil fuels and CO₂ emissions. Heat recovery in WWTPs, especially from treated effluent connected to district networks, offers very high technical potential (tens of TWh per year on a national scale in some countries) and can contribute significantly to more sustainable urban energy systems. Heat recovery in WWTPs can minimize the thermal impact of effluents on receiving rivers, reducing the negative effects of discharges on the natural environment.

1. Introduction

In December 2018, wastewater was officially recognized by the European Union as a renewable energy source, and heat recovery from it can be included in efforts to reduce greenhouse gas emissions [1]. Given that Wastewater Treatment Plants (WWTPs) are able to generate enormous heat excess, there is a margin to couple wastewater infrastructure to increase energy efficiency at the system level, to allow energy solutions to heat, to integrate volatile renewable electricity and, thus, to promote a sustainable energetic transition and a cleaner production [2].

The main relevant flows suitable for thermal recovery in a wastewater treatment plant can be identified, and each has distinct physical and energy characteristics, which determine the following recovery strategies: energy can be recovered from the raw influent, biological reactors, treated effluent, digestion reject streams and sludge (both raw and digested). However, in this study we are interested in the energy at the end of the process, as this is what can directly alter the temperature of the receiving river.

The temperature range at which anaerobic digestion occurs in the treatment process affects the recoverable energy, with a distinction being made between mesophilic digestion (≈35–37 °C) and thermophilic digestion (typically ≈55 °C) [3]. Although the thermophilic process ‘consumes’ more heat to operate at high temperatures, it also ‘produces’ more recoverable energy; in many cases, the net energy balance can be comparable or even favourable if the biogas and heat from the digestate are utilised effectively. In the context of our discussion, the effluent typically has temperatures higher than the ambient temperature because of the release of biological heat and constitutes a preferred source for recovery via heat pumps and other technologies; however, the exact temperature ranges of the effluent vary widely and depend heavily on climate, plant design and load [4,5].

Although some experience has already been reported in this area [6] the use of heat from wastewater is a relatively new concept in the field of renewable energy [7,8]. It could be said that the difference in temperature between a wastewater effluent―however highly treated it may be―and the receiving river is telling us about wasted heat energy that has also become a pollutant.

In general, when discussing heat recovery from water with excess heat, the focus is purely on energy utilization. Some analyse the possibility of heat recovery from residential buildings to WWTPs [9] from both technical and economic perspectives. Others focus on WWTPs from the same energy and economic perspective [10,11,12,13,14,15]. Finally, other case studies address the impacts of thermal discharges from WWTPs on receiving rivers [16,17,18,19,20,21], but we are not aware of any review that focuses on heat recovery at WWTP outlets and the use of this technology to reduce the environmental impact on receiving rivers. Kordana-Obuch et al. [22] recently conducted a bibliometric review on heat recovery throughout wastewater systems, and we have not found anything in their review that is similar to our approach. In our review, we aim to integrate this use into the context of minimizing environmental impacts, in which the use of thermal energy also serves to minimize the thermal impact of discharge from wastewater treatment plants, without ignoring other impacts that this type of discharge has on water quality.

A methodological section (Appendix A) has been included at the end of the document, describing how the literature search was conducted.

2. Importance of Temperature as an Ecological and Metabolic Factor

2.1. Temperature and Ecological Niche

Temperature is among the most significant environmental factors in relation to the physical, chemical, and biological processes that occur in inland waters [23,24] and, consequently, has a major influence on the biological success of fish and other aquatic organisms [25,26,27]. A general overview of the effects of warming on rivers, including an analysis of their ecological response and loss of resilience, was presented by Johnson et al. [28]. The relative importance of thermal emission impacts was found to be more relevant for the aquatic ecosystems of the Aare and Rhine Rivers than other stressors, such as chemicals and nutrients [29].

Under natural conditions, the temperature of continental waters ranges from −2 °C to almost 100 °C, and no organism can live across the entire range. In the Earth’s temperate zone, the temperature of continental waters normally varies between 0 °C and 25 °C, reaching 30 °C in tropical rivers [30] and 40 °C in rivers in hot desert areas [31]. The temperatures above occur only naturally in volcanic waters and hot springs [32]. Most aquatic organisms have little physiological control over their body temperature (they are poikilothermic organisms), so their metabolism is conditioned by the temperature of the water they inhabit. Some aquatic species can only live within a narrow range of temperature fluctuations (stenotherms), whereas others have a greater tolerance to thermal fluctuations (eurytherms). Most aquatic organisms live between 0 °C and 37 °C, although others can thrive in colder (−2 °C, marine notothenioid fish [33]) and warmer waters, such as some fish (e.g., Cyprinodon julimes, which can withstand temperatures of 44 °C and, eventually, up to 46 °C [34], among others [35]), and invertebrates (e.g., Thermosphaeroma subequalum, which can withstand temperatures of 45 °C [36]). Above 60 °C, only prokaryotic organisms are present, and only cyanobacteria can carry out aerobic photosynthesis under these conditions, with 73 °C being the maximum limit for any type of photosynthesis [37]. Above that temperature, only chemosynthetic organisms can thrive. Numerous thermophilic bacteria that inhabit hot springs such as those in Yellowstone have optimum temperatures in the range of 65 to 75 °C. In marine environments, where pressure allows for a high boiling point, some hyperthermophilic bacteria living in extreme environments have been isolated above 110 °C [37,38].

2.2. Water Treatment Temperature

Heat increases the activity rates of most chemical and physical processes, and organismal respiration increases with temperature within the temperature tolerance range of an organism [39]. Temperature can influence biological reactions in two ways: by influencing the rates of enzyme-catalysed reactions and by affecting the rate of substrate diffusion into cells [40]. Within the thermal range of microorganism function, most reaction rate coefficients increase as temperature increases, but eventually decrease as heat begins to inactivate cellular enzymes and denature other critical proteins and cellular structures within cells [38].

The cellular retention time, hydraulic residence time, redox conditions, temperature, concentration and acclimatization of the sludge used, as well as the technology employed, are common operating parameters in a WWTP and can significantly influence the treatment process [41].

Wastewater from homes arrives at a higher temperature than would be expected without the influence of domestic use, and this favours its biological treatment. In fact, reducing the temperature of the water in wastewater collectors could negatively affect its treatment or, at the very least, slow the work of the WWTPs [42]. However, WWTPs can release heat (e.g., nitrification and denitrification processes are exothermic). An example of the contribution to the heat balance of different elements in a WWTP in Rotterdam, Netherlands, is shown in Figure 1 [43].

3. Water Temperature and Climate Change

Many studies predict the impact of climate change on the temperature of inland waters [44,45] and, by extension, on the organisms that inhabit them [46,47,48,49,50].

Air temperature has tended to increase worldwide over the past century, and has become more pronounced since the early 1970s [51,52]. The IPCC’s Sixth Assessment Report (AR6) [52] predicts that widespread air temperature increases will continue throughout the 21st century. Using the IPCC Fourth Assessment Report (AR) SRES A2 scenario [53,54], Van Vliet et al. [45] predict that this increase will affect river temperatures by between 0.8 °C and 1.6 °C for the period 2071–2100 relative to the period 1971–2000. These changes are expected to be larger under the most pessimistic scenarios of the Sixth Assessment Report. They predict that the greatest increases will occur in the United States of America, Europe, eastern China, and parts of southern Africa and Australia.

Groundwater temperature is also affected by the temperature increases predicted by the IPCC models. Under a medium-severity scenario, groundwater at the depth of the water table (excluding permafrost) is estimated to have warmed by an average of 2.1 °C between 2000 and 2100, with the lowest warming rates occurring in mountainous regions such as the Andes and the Rocky Mountains [55].

Because of temperature increases, aquatic ecosystems have already experienced global water temperature changes (e.g., Australia [56], North America [57], Europe [58], and Asia [59]). This warming is exacerbated by reduced flow in many rivers, as has been shown by several authors [60,61]. Rivers with low base flows are more sensitive to warming [62]. Consequently, climate change will increase the impacts on aquatic ecosystems [63,64,65,66,67] and on habitat availability for many fish species because of increased temperatures and reduced flow rates [68,69,70]. Van Vliet et al. [45] predicted mean flow reductions in the range of 25–50% at southern and central latitudes in Europe (at the top of the range in the south, and at the bottom of the range at central latitudes, including the British Isles) and increases between 0% and 25% in the Scandinavian Peninsula, which could increase to 50% inland. However, it should not be forgotten that the climate scenarios used by these authors are more benign than the most recent ones from the IPCC Sixth Assessment Report [52].

4. Thermal Pollution

Caissie [24] defines thermal pollution as a reduction in water quality caused by temperature changes in natural water bodies as a result of anthropogenic activities. The causes of anthropogenic temperature alteration in natural water bodies can be grouped into: (i) deforestation and afforestation, (ii) damming of running water, and (iii) effluent discharge, including cooling systems for energy and industrial plants.

Deforestation has been observed to cause extreme elevations in water temperature in diverse biomes: temperature increases of up to 6 °C have been observed in the Amazon Basin [71] and 7 °C in Oregon (USA) [72], among many others (see the review of Moore & col. [73]). On the other hand, the afforestation of basins and riverbanks can induce temperature reductions in running water [74,75]. Damming implies the accumulation of water in reservoirs that reduces night–day oscillations because of the strong thermal inertia of the large, accumulated water mass (Figure 2). The mass of water influences its own thermal stability, as heat capacity is a function of specific heat and mass [76]. This means that a larger body of water (a river or a larger lake) is more thermally stable than a smaller one. Below 30 °C (the approximate temperature at which specific heat is at its minimum), the lower the temperature, the greater the specific heat and, consequently, for the same mass, the greater the heat capacity. Logically, these effects are in turn influenced by the ratio of the increase in surface area exposed to the atmosphere relative to the increase in mass [77]. A direct consequence of reservoirs is an alteration in the discharge temperature of the drainage water, which can undergo very significant cooling and heating with respect to the natural thermal regime, causing disturbances in the biological communities located downstream of the discharge [78,79,80,81,82,83]. Direct thermal pollution can occur when the water mass is used to cool a system (factories, power plants [84,85,86]) or as a receptor of warmer wastewater effluent [87].

Table 1. Alteration of the temperature with respect to that of the receiving mass of the outflow.

Stream/Basin (Season)	Distance to Outflow (km)	Temperature Increment (°C)	Source
South Plate River (winter)	0	7–15	[19]
	0.3	5–10
	5	3
St Vain (winter)	0	5–12	[20]
	0.3	7
	0.5	2
	9	3
Big Thompson (winter)	0	8	[20]
	6	4
Wissahickon (winter)	0	2–7	[17]
	>12	Effect
Wissahickon (summer)	0	1	[17]
	>12	Effect
Ner (winter)	effect on the duration of the ice season		[18]
Lower Rouge River (winter)	0	14.3	[16]
	5	9.9
	11	5.7
	19	4.1

summarizes some studies showing the longitudinal effect of thermal discharge on the temperature of receiving rivers. Urban stormwater is another factor of thermal pollution―that is assimilable to thermal discharge, as it is typically warmer than expected because of its circulation through the city [88,89]. With a lower density than the receiving water body, the contribution of heated water discharged into another water body causes density currents, which are particularly noticeable at the free surface and in the upper layers of the water, where it gives rise to a thermal plume [90]. After the heated water from the effluent and the river water has completely mixed, the change in water temperature depends solely on heat exchange with the environment. Heat exchange with the riverbed and banks can also be considered [90], which is negligible in most cases, as it is very small compared with the value of the heat exchange between the water surface and the atmosphere. However, when the heated water effluent is discharged near or on the riverbank, it can have a significant value [91].

Temperature alterations can strongly influence aquatic ecosystems, since most aquatic organisms tolerate only a relatively narrow temperature range (e.g., [92,93,94,95]). The consequences of warming on aquatic species range from minimal effects to lethal effects [29,96], although acclimatisation can occur, allowing animals to survive at temperatures somewhat higher than they normally would. However, aquatic plants are more resistant to high temperatures [84]. Not only the value of the recorded temperature but also the moment of the phenological cycle in which it occurs, is important since its fundamental phases can be altered (such as spawning [97] or hatching [46]). Direct effects include the aforementioned metabolic and reproductive alterations and, under extreme conditions, death due to damage to the nervous system [98,99].

5. Observed and Potential Thermal Impact of WWTP Effluents on Receiving Rivers

WWTP discharge alters the natural conditions of receiving water bodies, affecting their physicochemical and biological characteristics and, consequently, the metabolism of the receiving environment [100]. The thermal impact on receiving rivers can be significant and varies seasonally [21]. Wilson & Worrall [11] reported that, on average, wastewater effluent was 2.2 °C warmer than receiving river water, which can increase the temperature of the river. Such a thermal difference was also related to secondary and tertiary treatments in the WWTPs: compared with the Secondary Biological treatment (SB), Secondary Activated Sludge treatment (SAS) has a negative effect on stream temperature. On the other hand, compared with the Tertiary Biological treatment (TB), Tertiary Activated Sludge treatment (TAS) has a positive effect on stream temperature. The thermal influence of a discharge can range from a few hundred metres to several kilometres downstream, depending on factors such as the temperature difference, the ratio between the discharged flow and the receptor, the flow velocity, the turbulence of the mixture, the morphology of the channel, and environmental conditions (ambient temperature, solar radiation, wind).

Sewage effluent from WWTPs resulted in elevated metabolic rates in receiving fluvial systems. Zhang & Chadwick [101] reported that the eutrophication caused by WWPT is promoted jointly by nutrient supply and water temperature. When assessing the effects of alterations in the thermal regime, it is more useful to know the temperatures that organisms prefer and those they avoid than the temperatures that are lethal to them. However, both critical and lethal temperatures are only indicative of thermal impacts on aquatic ecosystems, since these impacts generally take place in streams with fluctuating temperature regimes, in contrast to the constant temperatures at which the tests for their assessment are conducted. Furthermore, in this case, the indirect effects can be far more significant than those directly caused by changes in temperature regimes that alter the aquatic environment and modify both the physical and biological habitats. Thus, in the first case, the alteration of water viscosity must be highlighted, which causes changes in the river substrate (an essential element of the benthic and interstitial habitats that are required by most aquatic species), as it affects the transport–sedimentation balance of solids. Temperature also controls the evaporation rate, and the saturation point of dissolved gases and solids, thereby influencing the concentration of vital elements and substances (such as oxygen and nutrients) and/or toxins in the water. High water temperatures may alter the toxicity of nutrients such as ammonia [102] and pesticides [103,104]. When an oxygen deficit occurs in water, the maximum lethal temperatures of several fish species decrease [105]. The dissolved oxygen concentration of a vegetated river is strongly related to its thermal regime and flow conditions [106]. In general, in polluted waters, an increase in temperature reduces the survival period of fish exposed to lethal doses. The indirect effects on the biological component can also be significant, as they affect all the trophic levels of the ecosystem. At each level, the metabolism, growth, and reproduction of populations are altered, which in turn affects all their interactions (competition, predation, and parasitism), which are frequently the regulatory mechanisms of the entire ecosystem.

5.1. Effects on Primary Producers

Primary producers in rivers are represented by periphyton, phytoplankton, and macrophytes. Increased water temperature generally corresponds to an increase in the biomass of these organisms and in primary production. Maximum algal growth occurs at different temperatures, depending on the species [107,108]. These maximum growth rates can generally be summarized as follows [109] (

Table 2. Temperature range of maximum growth [109].

Group	Temperature
Diatoms	15–25 °C
Green algae	25–35 °C
Cyanobacteria	30–40 °C

):

Therefore, a temperature increase generally leads to the disappearance of the most demanding cold-water species and their replacement in the community, or an increase in abundance, by more thermophilic species. Above 30–32 °C, any further temperature increase led to a decrease in algal diversity, resulting in a tendency for the community to be dominated by cyanobacteria.

Studies on thermal pollution produced by power plants have shown that moderate temperature increases (2–5 °C) generally stimulate primary production between spring and autumn [110,111], although more pronounced increases or high temperatures in summer often result in photosynthetic inhibition and phytoplankton mortality [110]. The underlying mechanisms include direct thermal effects on metabolic rates and photosynthetic efficiency, as well as indirect effects mediated by changes in thermal stratification, nutrient availability, herbivory pressure, and dissolved oxygen budgets. The average annual primary production in a thermal power plant whose temperature increased by an average of 6.3 °C resulted in a 27% increase in annual primary production [112]. However, aerobic and anaerobic processes, as well as the balance between respiration and production, depend not only on the effluent but also on the characteristics of the receiving river.

The temperature increase induced by WWTP discharge, which is particularly pronounced during winter [20], can amplify the effects of nutrient enrichment by accelerating metabolic rates. However, studies suggest that compared with primary production, warming increases respiration more rapidly [113]. In WWTP discharges, the interaction between elevated temperature and nutrient load can intensify the proliferation of phytoplankton and benthic algae, alter species composition by favouring cyanobacteria and tolerant taxa, and shift the metabolic balance of the ecosystem toward heterotrophic conditions [114,115]. Yang et al. [21] reported increases of 84% in phosphate and 19% in nitrate in receiving rivers, which can translate into substantial increases in primary production when other factors (light, temperature, residence time) are favourable. However, this subsidy effect can be modulated or counteracted by stress factors associated with discharge, including trace contaminant toxicity, high turbidity, or inhibitory concentrations of ammonium [116].

The effects on the species composition of phytoplankton communities are equally important as the effects on total biomass. WWTP discharge favours species that are tolerant to eutrophication and degraded water quality conditions [117] and may promote the development of potentially toxic cyanobacteria [118]. These changes in species composition have implications for water quality, trophic network structure, and the provision of ecosystem services. Furthermore, the level of wastewater treatment is critical in determining the magnitude of the impacts. Tertiary treatment, which includes specific nutrient removal processes, is significantly more effective than secondary treatment at reducing nitrogen and phosphorus loads [21]. However, even with tertiary treatment, effluents contain residual amounts of nutrients and organic matter that can have detectable ecological effects, especially in rivers with low dilution capacity.

Nevertheless, the dilution capacity of the receiving river, which is determined by its flow rate, emerges as the main factor modulating the effects of discharge. The effects are typically more pronounced during periods of low flow, when the dilution capacity is reduced and the hydraulic residence time is prolonged [119]. It is therefore important to know the annual flow regimes of both the discharger and the receiver. Dilution of both, nutrients and heat excess, is important, where applicable, as the combined effect of these two factors increases the metabolism of the receiving environment and, ultimately, affects the aquatic community both qualitatively and quantitatively.

5.2. Effects on Macroinvertebrates

Water temperature is among the most fundamental abiotic factors that structure macroinvertebrate communities in lotic ecosystems, influencing their physiological processes, metabolic rates, ontogenetic development, behaviour, and survival [120], and these communities are particularly sensitive to temperature variations because of their ectothermic nature and life cycles, which are closely synchronized with seasonal temperature regimes [107,121,122,123,124]. In addition, rising temperatures can amplify the toxic effects of chemical pollutants on aquatic macroinvertebrates, as has been demonstrated in numerous studies (e.g., [104,125,126,127]). This relationship between temperature and the biology and ecology of macroinvertebrates has been extensively studied because of the potential effects of global warming on their populations and communities; thus, making inferences about the effects of thermal discharge from these studies is primarily a question of scale.

The growth of macroinvertebrate species (especially the most thermophilic ones) increases with increasing temperature, provided that there is sufficient food in the aquatic environment. There is clear evidence that macroinvertebrates develop higher feeding rates at higher temperatures, but they may experience reduced body size at maturity, altered reproductive phenology, and even an unbalanced sex ratio [128]. If the temperature increase is significant, feeding activity may cease long before lethal temperatures are reached, thus halting growth. Thus, although rising temperatures increase macrobenthic growth, their reproductive potential is reduced by the production of relatively small adults.

Macroinvertebrates can be classified according to their thermal preferences into categories ranging from cold stenotherms (narrow range, low temperatures) to eurytherms (wide range) [129]. At the community level, warming can lead to species replacement, with psychrophilic taxa being displaced by thermophilic species [130,131]. Thermal tolerance studies have identified families that are particularly sensitive to warming, including Paramelitidae, Notonemouridae, Teloganodidae and Philopotamidae, as well as moderately sensitive families such as Palaemonidae, Heptageniidae, Leptophlebiidae, Corydalidae and Aeshnidae [120]. The maximum water temperature explains up to 29% of the variability in the Ephemeroptera, Plecoptera, Trichoptera and Coleoptera communities in small and medium-sized rivers [124]. In the United Kingdom, an analysis of more than 2300 rivers over 21 years revealed that long-term changes in taxon prevalence correlated with sensitivity to pollution and flow, although short-term (<2 years) thermal effects were evident [132]. Worthington et al. [133] reported that in the Severn River, the abundance and richness of the macroinvertebrate community decreased more than 0.5 km downstream below a discharge of 4.5 °C above ambient temperature. The abundances of Musculium lacustre, Simulium reptans, and Orthocladiinae were greater at the unheated control site, whereas more pollution-tolerant species such as Asellus aquaticus and Erpobdella octoculata were more abundant in the thermally impacted reaches. In parallel to the decline in psychrophilic species, warming facilitates the emergence and colonization of limnophilic and thermophilic taxa. In the Loire River, the expansion of invasive species such as Corbicula sp., an Asian bivalve that is tolerant to high temperatures, has been observed [130]. Pander et al. [134] specifically examined the effects of thermal variability on native versus nonnative gammarid species and reported differential responses that favour nonnative species under altered thermal regimes. In lowland rivers with hydromorphological degradation, the increase in temperature contributed to a 150% increase in the total abundance of macroinvertebrates, as well as in the abundance and richness of sediment-dwelling and nonnative taxa, in which case the increase was as high as 235% [135].

The effects of warming on taxonomic richness are complex and context dependent. In the thermal manipulation experiment by Hogg et al. [128], no significant changes in taxonomic richness were detected despite reductions in density. However, predictive models for the Spanish Mediterranean area indicate drastic reductions in taxon richness under scenarios of thermal increase, with 65.96% of taxa inhabiting medium-high elevations experiencing contractions in their distribution area [136]. On the other hand, in Finland, Mustonen et al. [137] modelled increases in taxonomic richness, particularly in the north, which is consistent with northwards shifts in the distributions of many species and the replacement of others. Other examples of structural changes in communities can be found, for example, in Li et al. [138] and Fruget et al. [139]. All these changes can give rise to other functional changes, affecting the base of aquatic food webs, with cascading consequences for consumers, including macroinvertebrates. Rising temperatures and increased primary production, mainly in the form of phytoplankton, will lead to functional changes in the macroinvertebrate community: grazing and shredding organisms will give way to a community dominated by collectors, in an environment where the P:R (production versus respiration) ratio will be shifted towards respiration [140]. In this environment, invasive species may find conditions favourable for proliferation.

5.3. Effects on Fish

As in the case of macroinvertebrates, fish are ectothermic organisms in which the internal temperature and the temperature of the surrounding environment are closely linked, directly affecting metabolic processes and growth expectations [141,142,143,144]. In addition to growth inhibition, temperature increases can cause pathological crises in fish (susceptibility to contaminants [145] and infections [146,147,148]). There is a direct, curvilinear relationship between oxygen consumption at rest in fish and ambient temperature [149]; for example, a tropical fish at a temperature of 30 °C needs approximately six times more oxygen than a polar fish at 0 °C does [150]. Heat stress also triggers a cascade of physiological responses, including the activation of the hypothalamic-pituitary-adrenal axis, the production of heat shock proteins (HSPs), changes in oxidative metabolism, and alterations in ionic balance. These responses, although initially adaptive, can become harmful when stress is chronic or when the organism’s acclimatization capacity is exceeded [151,152]. Pregler et al. [153] reported that temperature variation can generate interspecific synchrony but spatial asynchrony in the survival of freshwater fish communities, with implications for the regional stability of populations. Moreover, the discharge of thermal water into cold water (or the release of cold water into warm bodies of water) attracts certain fish species, especially during autumn and winter, leading to concentrations and migrations that differ from natural patterns [123,154]. Réalis-Doyelle et al. [155] reported differential mortality depending on the development temperature of brown trout embryos and fry. Sessions et al. [156] reported that the heat shock response exhibits plasticity in lake whitefish embryos (Coregonus clupeaformis) exposed to repeated heat stress, suggesting acclimatization ability at the molecular level. Pandey [152] documented concurrent changes in thermal tolerance thresholds and cellular responses to thermal stress, revealing new molecular signatures and markers of acclimatization to high temperatures in rainbow trout. Through its impact on feeding efficiency and body condition, temperature can also affect competitive interactions in invasive and native freshwater benthivorous fish [157]. Coulter et al. [158] documented species-specific effects of subdiurnal temperature fluctuations on consumption, growth, and stress responses in two physiologically similar fish species. Other physiological functions that may be affected by temperature include protein metabolism [138,159]; osmoregulation [160]; the timing of reproduction [161]; cardiac function [162,163]; gene expression [164,165]; blood and reproductive maturation [166]; fish behaviour [167,168]; and other haematological, physiological, gene expression and behavioural variables [169]. Lukšienė et al. [170] reported that high temperature thermal effluent areas negatively influence the gametogenesis of the female perch Perca fluviatilis, roach Rutilus rutilus and pike Esox lucius, resulting in reduced reproductive capacity. In experiments with brown trout, warming increased damage to sperm and reduced oocyte fertility, as well as causing asynchrony in maturation in the most extreme scenario [171].

Few fish species inhabit water with temperatures exceeding 30 °C. Stenothermic species do not tolerate oscillations greater than 8–10 °C. In temperate waters, most species are eurythermal, although their optimal temperature range varies with the season and stage of development [172]. There are also lethal temperatures, which as an example are set out below, such as the maximum temperatures of some species (

Table 3. Lethal temperatures of several fish species [105].

Specie	Temperature
Rainbow trout, Oncorhynchus mykiss	25–28 °C
Brown trout, Samo trutta	23–30 °C
Atlantic salmon, Salmo salar	28–30 °C
Pike, Esox lucius	28–34 °C
Sea lamprey, Petromyzon marinus	34 °C
Common Sunfish, Lepomis gibbosus	34 °C
Black-bass, Micropterus salmoides	32–36 °C
Gudgeon, Gobio gobio	36 °C
Tench, Tinca tinca	29–39 °C
Goldfish, Carassius auratus	31–38 °C
Carp, Cyprinus carpio	31–40 °C

), although this also depends on the populations, water quality, exposure time and the acclimatisation temperature:

In summary, at the individual level, warming reduces thermal safety margins, increases physiological stress and can cause direct mortality when critical thresholds are exceeded. At the population and community level, warming can cause changes in the distribution, composition and structure of fish communities. Physiological and even behavioural acclimatisation may help fish to survive, but the decline in their competitive ability―linked to a reduction in their functional capacity (such as their ability to swim away from predators or to feed [173])―can become an insurmountable disadvantage when faced with competitors better adapted to the new conditions, which would ultimately alter the community.

5.4. Effects of Warming on River Communities

The response of a river’s biological community to rising water temperatures depends largely on its composition. If species are highly sensitive to thermal changes, and if they have a stenotic thermal spectrum, their impact will be much greater. Species adapted to cold waters (cryophilic) are replaced by thermophilic species as rivers warm up, becoming confined to colder headwater sections until they may eventually disappear. However, some species may benefit from warming including invasive species (e.g., [174,175,176]).

In general, compared with communities with low diversity, well-structured communities with high diversity are more strongly affected. In mountain rivers, water temperature and flow regimes are strongly linked to climatic seasonality, and therefore naturally fluctuate considerably. Conversely, in the lower reaches of rivers, the water originates from a relatively large area where the heterogeneous characteristics of the headwaters and tributaries mix, resulting in much smoother temperature and flow regimes downstream. River communities, throughout their evolution, have adapted their life strategy to these types of fluctuations, and therefore are distributed along the river following a continuous altitudinal gradient, or zonation [177], whose different compositions, structures and functions indicate different responses to different environmental conditions [140]. The effects of temperature can range from very local to very widespread, and the severity of their consequences for the receiving system will depend on the specific circumstances.

6. Wastewater Temperature Recovery

Thermal energy can be recovered with high efficiency from the source in the buildings where it is generated [178] or in wastewater treatment plants [179]. The potential for recoverable thermal energy from wastewater is much greater than the chemical energy contained in it [180]. In general, wastewater treatment involves increasing the water temperature to increase biodegradation. Ideally, the energy from treated wastewater can be used in the WWTP itself, e.g., for heating the digester tank or for low-temperature sludge drying. Both applications allow the energy of wastewater to be harnessed at a temperature level that is attractive for heat pumps [181].

Wastewater effluents are typically warmer than the rivers to which they are discharged. Excess thermal energy could be recovered, minimizing the thermal impact on the receiving river. Wilson and Worral’s [11] analysis using data from England for the period 2000–2019 revealed that effluent temperatures were on average 2.2 °C higher than receiving river temperatures, with a corresponding annual heat recovery potential of ∼18.3 TW h that could meet ∼3.6% of the UK heat demand. The temperature of the raw wastewater (wastewater that has completed treatment before leaving the facility) was on average 1.5 °C higher than the effluent temperature, implying that an additional ∼12.5 TW h is lost annually during treatment before discharge. In the same study, the greatest difference in temperature between the effluent and rivers and between the raw wastewater and effluent occurred during the autumn and winter months, indicating that the highest seasonal heat recovery potential coincides with the highest heat demand. The temperature difference between the effluents and rivers increased at an average rate of ~0.03 °C per year between 2000 and 2019. Wastewater treatment plants (WWTPs) discharge continuously and are located near human settlements, which account for most of the heat demand. Therefore, heat can be recovered and, in addition, the environmental impact on rivers can be reduced. Furthermore, Neugebauer et al. [10] calculated that in Austria alone, wastewater treatment plants can provide 40% of the installed thermal capacity in district heating networks without resorting to combined energy sources.

There are already numerous examples of heat recovery from wastewater in Switzerland, Finland, Norway, Sweden, Canada, Japan, and so forth [11]. The first WWTP with a heat recovery system was in Obermeilen (Switzerland) and was built in 1975 [182].

The first heat recovery plant in a Japanese WWTP was built in Tokyo in 1987 to serve its administrative buildings [183]. The authors report that the heat energy lost in the sewage system of the Tokyo metropolitan district is equivalent to the energy consumed by 400,000 homes for heating and cooling, which suggests the potential for energy recovery. Another major project was carried out in Sapporo, where recovered heat energy was reused to melt snow on the streets [183]. Mikkonen et al. [184] estimated the recoverable heat energy from wastewater in Sweden at 720 GW h per year for each degree of temperature recovery. The use of heat pump technology could also increase the overall efficiency of the heat recovery system by providing cooling energy during the summer season [178]. Lingsten and Lundkvist [185] quantify the consumption of water and sewage services in Sweden at 1.3 TW h of electricity and approximately 0.5 TW h of other energy sources, excluding the energy content of the chemicals used in water treatment. WWTPs are the largest electricity consumer at 630 GW h, with aeration consuming 24% of the electricity. In turn, the WWTPs produce approximately 0.6 TW h of biogas and approximately 2.5 TW h of thermal energy. Đurđević et al. [186] modelled wastewater utilization in the city of Rijeka, Croatia. The case study site had an operating WWTP with a capacity of 540,000 population equivalents and 3000 L/s of effluent water at full load. On the basis of the considered water flow rate and a temperature decrease of 6.5 °C, a heat recovery potential of 75 MW was obtained, which was 72% of the capacity of the existing natural gas plant.

The energy potential of treated wastewater is much greater than that of raw wastewater. This is because, downstream of the wastewater treatment plant, the wastewater can be cooled much more than upstream―up to 8 °C. For aquatic fauna, such cooling of the wastewater is even desirable. Unfortunately, the high energy potential of treated wastewater cannot be utilized in many locations because wastewater treatment plants are located outside residential areas, where heating is unavailable [181].

Ideally, the energy from treated wastewater can be used within the WWTP itself, either for heating the digester tank or for low-temperature sludge drying. Both applications allow the energy of wastewater to be harnessed at a temperature level that is attractive to heat pumps. However, there are few examples of the internal use of wastewater heat in WWTPs. This is because many wastewater treatment plants have large amounts of waste heat available from the use of waste gases in combined heat and power units. In the future, this idea could attract increasing interest if larger wastewater treatment plants increasingly condition their waste gas to meet natural gas quality standards and thus be able to feed it into the public gas grid [181].

In Switzerland, in 2008, 20 WWTPs externally utilized heat from treated wastewater [181]. A distinction can be made between two heat supply systems: cold and hot district heating. In the former, treated wastewater is extracted from the wastewater treatment plant outlet and pumped through a “cold” main pipeline to the consumer. Heat generation using heat pumps occurs decentrally. After heat extraction, the cooled wastewater is returned to the wastewater treatment plant or discharged directly into a receiving stream or river. In the case of the “hot” district heating system, usable heat is generated centrally at the wastewater treatment plant or in a neighbouring building. In the case of the Swiss capital, the wastewater treatment plant is designed for approximately 350,000 residents. It has a heat recovery potential of more than 30 MW from treated wastewater. Part of this potential (1400 kW) is used in the heating system of the neighbouring Bremgarten district. The Bremgarten Cooperative sells a total of 5 GWh of heat per year. Approximately 60% of this heat comes from wastewater. Owing to the significant difference in height between the wastewater treatment plant and heat users, the network is divided into three sections with intermediate heat exchangers. The measured annual coefficient of performance of the wastewater heat pump system (including network pumps) is 3.0 [181].

Other notable examples include Mokravica (Poland), Gaobeidian (China) and As-Samra (Jordan). In the case of Mokrawica, recovery estimates from various points within the WWTP showed average power outputs of 309–451 kW and winter minimums of 58–68 kW at certain locations within the plant [14]. With respect to the influence of climate, the analysis of recoverable power by plant location quantified average and winter minimum ranges, which is useful for designing heat pump recovery systems [14]. At the Gaobeidian WWTP, a full-scale case is illustrated in which the energy recovery system already in place achieved 268,788 MWh/year against a consumption of 280,717 MWh/year, and it was estimated that an increase in the recovery of waste heat from the effluent could increase the energy neutralisation ratio to over 500% in extreme scenarios. This is an example of a plant that has almost achieved energy neutrality because of integrated energy recovery and utilisation; the results of this study demonstrate the synergistic effect of waste heat recovery and biogas on the overall balance [187]. The As-Samra case is of particular interest because it concerns a plant located in a hot climate [188]. The ORC integration analysis and economic evaluation revealed that an optimal ORC installation of ~412 kWe was viable with specific NPV and LCOE figures for that context, illustrating how technological implementation and energy analysis depend on the local operational profile. The main sources of waste heat identified at the As-Samra WWTP are the exhaust gases from the combustion process and the circulating water in the generator jackets. This heat can be utilised in various applications, such as power generation, cooling and water desalination. The implementation of WHR systems in WWTPs is crucial because of the significant amount of heat generated during the treatment process. If this heat is not recovered and then utilised, it is lost to the environment, resulting in the inefficient use of energy resources. Heat recovery can improve overall energy efficiency, reduce reliance on external energy sources and decrease the environmental footprint by minimising waste heat emissions. Recovered heat can be used to maintain optimal temperatures in anaerobic sludge digestion, provide heat for space heating within WWTP facilities, preheat water for various processes, and drive systems such as the Organic Rankine Cycle (ORC) to generate additional electricity [188].

To determine the thermal recovery potential at a specific WWTP, it is essential to have local measurements of temperatures, flow rates and regulatory constraints; the cases cited demonstrate processes and magnitudes but do not allow for the generalisation of universal thermal ranges without local data [13,14,187,189,190].

7. The Normative Matter

From a regulatory point of view, when the case of the European Union is analysed, the use of waste heat from WWTP effluents has not yet been given much consideration. RED III [1] places the recovery of thermal energy from WWTP effluent in a grey area. This Directive distinguishes between “waste heat”, which is recognized only when it is discharged into a district heating network, and “ambient energy”, which explicitly includes the heat contained in surface water and sewage water, allowing thermal energy extracted from wastewater already integrated into the sanitation system to be classified as renewable. As Holzleitner-Senck et al. reported [191], this distinction leads to paradoxical results: the direct and efficient recovery of heat from WWTP effluent using a dedicated circuit and heat pumps―without passing through a district network―may fall outside the legal categories of waste heat and renewable energy, while the recovery of the same thermal flow a few metres downstream, already in the collector, could be considered renewable ambient energy for the purposes of RED III targets. This would allow heat recovery from treated water to fall within the definition of waste heat in RED III but more as a kind of semantic loophole than as a matter of substance. Consequently, consideration should be given to revising the definition of waste heat in RED III to explicitly include the use of heat from treated wastewater, regardless of whether it is used in district networks or direct supply schemes, in line with the principles of energy efficiency and decarbonization that inspire both RED III and the Energy Efficiency Directive. In the case of the U.S., we have not found a federal definition of waste heat applicable to WWTPs, although incentives under the United States Code may be applicable. 42 U.S.C. § 6343 [192].

The examples shown make it clear that heat recovery is technically and economically viable. However, this is an area that still needs to be explored in greater depth, as few steps have been taken thus far and more proactivity is required on the part of the agents involved (energy and environmental agencies, water managers). An appropriate regulatory framework must be created that goes beyond political intentions alone (as is the case with other sources of self-generated renewable energy). Economic viability must be studied on a case-by-case basis, as in many cases, it is a question of scale (affordable in large facilities where recovery is significant), or in certain cases, this option must be treated strictly as an environmental correction measure regardless of its economic profitability.

8. Final Discussion

Experience has shown that wastewater heat pumps operate efficiently. Primary energy consumption relative to the useful energy produced is much lower than that of traditional heating and cooling systems. Compared with a condensing gas heater, a wastewater heat pump (with a peak-load boiler) uses 10% less primary energy, and 23% less primary energy than an oil heater. Furthermore, compared with other heat pump systems (groundwater and geothermal probes), wastewater systems perform well. This is because the heat source has favourable temperatures year-round. When properly planned and optimally operated, wastewater systems achieve high annual coefficients of performance. The highest value measured in Switzerland at a facility in Basel is greater than 7 [181], and the CO₂ footprint is reduced by 78% compared with that of an oil heating system.

Recovering heat energy from WWTP effluent has advantages over recovering wastewater closer to its source, before it enters the WWTP [8,193]: (i) flow rate and temperature variations are typically small and predictable, which improves system management [16]; (ii) the use of treated wastewater reduces equipment fouling and clogging; and (iii) it has no negative effects on biological wastewater treatment [194]. Moreover, although heat energy is lost on its way through the sewer network [195], the exothermic biological processes occurring in wastewater treatment plants also increase its temperature. Recovered heat can be used internally at WWTPs, but the amount often exceeds internal heat requirements, and it is feasible to deliver the excess heat to a local district heating system, if it exists.

In 2005, thermal energy utilization through heat pumps was estimated to produce 22% of the CO₂ emissions produced by an oil-fired plant and 35% of those produced by a gas-fired plant [196]. The efficiency of heat recovery technology is constantly evolving; thus, these figures are expected to improve. Furthermore, although further development is needed, thermal energy can be stored in thermal accumulators, using materials with suitable thermodynamic properties to capture the heat produced [197].

One issue of concern is the heat loss between the recovery station and the end user. Neugebauer et al. [10] highlighted the importance of the spatial proximity of heat recovery locations for potential users. Factors to consider include the number of operating hours and thermal losses due to heat transport [198]. These contributions have provided useful insights into the impact of temporal and spatial variations on the performance of heat recovery systems.

Spriet et al. [2] proposed a three-step methodology to ensure the optimal use of the thermal energy at hand, which includes an energy analysis at the wastewater treatment plant, a spatiotemporal analysis of supply and demand in areas of potential supply, and an integrated analysis, superimposing the supply and demand profiles. These authors concluded that wastewater constitutes a suitable energy source to supply baseloads, but the spatiotemporal patterns reveal that both periods of excess wastewater heating potential and periods requiring additional heating will occur in bivalent systems. Therefore, the urban and regional spatial grid, the mix of land uses and their density largely affect the design and usable quantity of this renewable energy source. Finally, the use of wastewater energy provides viable and valuable contributions to sustainable urban energy supply systems and cleaner production if the electricity sources for the respective heat pump systems are renewable and guarantee low- to zero-emission operation.

For obvious operational reasons, natural temperature fluctuations throughout the day cannot be accurately replicated in the effluent; therefore, working with a deviation of ±1 °C from the average daily temperature of the receiving body at the time of discharge seems, a priori, more than acceptable. It is also important to adjust heat production to actual heat demand, and together these considerations should promote the development of integrated models that jointly address energy production and impact minimisation. The possibility of generating automatic models that integrate real-time data on influent and effluent temperatures and energy requirements now seems feasible given the available technology and the processing capabilities of artificial intelligence.

To assess the suitability of a heat recovery strategy in wastewater treatment plants and ensure its correct implementation, we propose a procedure (Figure 3) that begins with an energy analysis of the treatment plants and a spatio-temporal analysis of energy supply and demand. This should be followed by an analysis of the impact of thermal discharges and a spatio-temporal assessment of the river system’s tolerance margins to temperature changes, culminating in an integrated analysis for the optimal use of thermal energy and the minimisation of environmental impacts. To fully exploit this potential, several lines of work must continue to be developed, including thermal energy storage, improving the efficiency of recovery technologies, refining the definition of tolerable margins of temperature variation in the receiving river, and strategic planning that explicitly takes into account the spatial proximity between sources and demands. Taken together, these guidelines can contribute to more efficient energy management of wastewater discharges and to the conservation of natural resources affected by effluents from wastewater treatment plants.

9. Conclusions

As a corollary, this document presents a set of general conclusions and proposals to advance the development of heat recovery as a means of reducing the environmental impact of WWTP discharge, particularly thermal impact. Water temperature has emerged as a key ecological and metabolic factor in rivers, since it influences species distribution, metabolism, growth, reproduction, and survival rates. Thermal pollution caused by human activities such as dams, discharges, urban stormwater, and industrial cooling alters the natural thermal regime of rivers, thereby modifying the structure and functioning of communities of primary producers, macroinvertebrates, and fish, and favouring thermophilic species. Wastewater treatment plants (WWTPs) generate and discharge excess heat, and their effluents are often several degrees warmer than those of the receiving river; this increases the metabolism of aquatic communities, favours eutrophication, and can intensify the effects of nutrients and toxic pollutants, as well as the replacement of species. Moreover, excess heat from wastewater represents a major renewable energy resource that can be recovered using heat pumps, both in buildings and in the treatment plants themselves, as well as in district heating networks, thereby reducing the demand for fossil fuels and associated CO₂ emissions. In particular, heat recovery in WWTPs, especially from treated effluent connected to district networks, offers very high technical potential and can contribute significantly to more sustainable urban energy systems, while also minimizing the thermal impact of effluents on receiving streams and reducing the negative effects of discharges on the natural river environment. Nevertheless, research on the range of temperatures tolerable by the receiving system, which is essential for defining suitable operating temperatures for energy recovery, is lacking. In this context, there is a need to develop models that incorporate river ecological requirements, particularly those relating to instantaneous temperature and energy demand, to adjust heat recovery. Models that combine real-time data with decision-making systems powered by AI could provide a means of achieving these objectives.