What is a Shallow Lake?
The ecology of lakes is strongly influenced by physical factors, particularly the depth of light penetration and the depth to which the water is mixed by the action of wind. Thus lake depth is very important in determining how a lake responds to environmental pressures.
A useful definition of a shallow lake is a lake where the water column is fully mixed and in the absence of phytoplankton growth sufficient light is able to reach the sediment surface to enable photosynthesis to occur. Thus in, good condition, the open water of a shallow lake is likely to be dominated by submerged macrophytes.
A useful operational definition is a saucer shaped basin with a mean depth of < 3m, although a larger slightly deeper lake (<5m) with a substantial fetch might also behave in a similar way. The difference between a shallow lake and a pond is essentially size, with water bodies <100m maximum diameter considered to be ponds due to increased prevalence of chemical stratification.
Why are Shallow Lakes important?
- Shallow lakes provide a rich habitat for a wider variety of aquatic plants, invertebrates and fish.
- Shallow lakes are very common, the UK lake inventory (Hughes et al. 2004) contains 8420 shallow and very shallow lakes (99% of all lakes) with 30% of these being very shallow (mean depth < 3 m).
Threats to Shallow Lakes
- Eutrophication. Enrichment by nutrients is probably the most important pressure influencing shallow lakes. Sources of nutrients to lakes are multiple and include sewerage effluent, septic tank discharges, agricultural run-off and excessive wildfowl/fish stocking among many others. The majority of shallow lakes are found in lowland areas, where agricultural and human activity results in elevated levels of nutrients, particularly phosphorus and nitrogen. Although concentrations of phosphorus in rivers have been falling since the 1990s, 74% of monitored lakes in England exceed the WFD standard for good status (Environment Agency 2014).
- Terrestrialization. Lakes are generally of sufficient size and depth to not be significantly threatened by terrestrialization. However, littoral habitats can become shaded by trees and where this is accompanied by the regression of the marginal vegetation due to other pressures (Boar et al. 1989) the littoral margin may be substantially reduced.
- Habitat Disturbance. Shallow lake reedswamps and marginal habitats can be severely damaged by a variety of disturbance factors; including boating activity (excessive boat wash), poorly-managed shoreline access and overgrazing from wildfowl.
- Toxicity. Because of their location and often more intensive use by human activities, contamination to toxic levels can occur due to effluent inputs. Mixing and recycling between contaminant loads in sediments can occur more easily due to shallowness. Negative ecological effects for aquatic ecosystems from metals, anti-fouling (TBT) paints, pesticides are measurable in many shallow lakes.
- Over-stocking with Fish. Shallow lakes used for angling can often suffer deleterious consequences (algal blooms, loss of aquatic plants) due to fish overstocking, especially true of common carp (Cyprinus carpio).
The Water Framework Directive
The tools developed to assess the status of lakes for the Water Framework Directive (WFD) are the most appropriate method of assessing the current status of shallow lakes (UK Technical Advisory Group). These methods include assessments of phytoplankton, macrophytes, phytobenthos and invertebrates, but at present they do not include fish. Each of these biological quality elements is placed into classes ranging from High to Bad ecological quality and can be used to provide a guide to the overall status of a shallow lake. It is important to recognise that while on average these methods provide a reliable assessment of lake status, they will not always reliably reflect conditions of any particular lake and results should be interpreted using local knowledge and wider information (Phillips, 2014).
- EU Water Framework Directive. This EU Directive is implemented by the Environment Agency and requires all lakes over 50ha to be restored to good ecological status by 2015. (Alter for different habitat types)
- EU Habitats Directive
Principles of Shallow Lake restoration
Importance of macrophytes
Macrophytes, large aquatic plants, are extremely important in shallow lakes offering habitat for other organisms (Carpenter and Lodge 1986) and crucially influence food web interactions in shallow lakes (Burks et al. 2006), which are so important in determining species communities (Scheffer et al. 2001). Macrophytes have been lost or greatly diminished in many shallow eutrophic lakes (Jeppesen et al. 2000; Sand-Jensen et al. 2000) and their re-establishment should be a key aim in shallow lake restoration (Bakker et al. 2012)
Work at the catchment/landscape scale
Lakes are a product of their catchment, so it is essential to identify the catchment and consider the pressures that arise from it. Although internal processes such as nutrient re-cycling and fish communities will have important influences on the lake, it is the external nutrient load that is most likely to determine long-term condition
Realistic time scales and expectations
Although ecological change can be relatively rapid, recovery from nutrient enrichment is a slow process. It is likely to depend on the rate of water exchange, but typically it is likely to take in excess of 15 years for substantial change to occur (Jeppesen et al. 2005b).
What are the expectations for the lake? If the aim is simply to reduce the incidence of nuisance algal blooms then this may be quicker and easier to achieve than re-establishing a diverse macrophyte community. However, in most cases a longer-term plan to restore the lake is more likely to lead to stable outcomes. For shallow lakes this almost certainly requires establishing a diverse macrophyte community that covers at least 30% and preferably 50% of the lake area (Bakker et al. 2012). In some cases this may be incompatible with current uses of the lake, for example sailing and perhaps fishing and it thus important to discuss and agree goals with a wider variety of stakeholders.
It is important to establish as much about the lake as possible when planning a restoration. What is the current status, what information is available about its past condition, what is known about fish stocks, what are the major sources of water? In most cases it is invaluable to be able to monitor the current status for a period of at least 3 years before management work is undertaken. For large scale and/or expensive restoration measures consideration of availability of macrophyte propagules (seeds or spores) should be undertaken.
Nutrient concentrations, especially P &N is the key element for assessing lake status. Soluble forms of nitrate and phosphate are easier to measure but can give misleading results as they are rapidly taken up by plants, particularly algae and thus underestimate nutrient availability. To overcome this total phosphorus and nitrogen should be measured on unfiltered samples as the chemical digestion process includes nutrients bound in phytoplankton. There is significant seasonal variation in nutrient concentration and it is important to take samples throughout the year to determine true status. Ideally 12 monthly samples should be used, but as a minimum 4 quarterly samples are required. If it is not possible to measure total nutrients winter concentrations may provide an indication of nutrient availability and thus trophic status. Results can be compared to standards established under the WFD.
It may not be possible to apply the full WFD assessment for phytoplankton. A simpler approach would be to determine the mean annual chlorophyll a concentration, using water samples collected to determine nutrient status. As for nutrients, samples must be collected throughout the year. The resulting mean chlorophyll a concentration can be used in the UK WFD phytoplankton assessment tool (PLUTO) to provide an indication of status (UK Technical Advisory Group 2014). The ratio of the mean chlorophyll a (µgl-1) to total phosphorus concentration (µgl-1) also provides an indication of the extent that grazing controls phytoplankton biomass, with values of…
Macrophytes, Zooplankton and Fish
A full understanding of lake status can be gathered by studying macrophyte communities and by recording populations of zooplankton and fish. Specialist ecologists can be brought in to generate this data.
Controlling nutrient loading
Reducing the nutrient load, particularly that of phosphorus, from the catchment is likely to be the single most important issue when considering the restoration of a shallow lake. Despite considerable investment in the reduction of phosphorus from larger sewage treatment works (STW) 45% of classified rivers in England and 7% in Wales fail the phosphorus concentration established under the WFD to support Good ecological status. For lakes the situation is worse, with 76% failing the standard in England and 36% in Wales (Environment Agency, 2012).
- Sources of phosphorus
The most important task is to identify sources of phosphorus in the lake catchment. The largest source (60-80%) of phosphorus to rivers is generally considered to be from STWs, whereas for lakes it is often reported to be agriculture (Environment Agency, 2012). This generalisation is partly a reflection that many smaller lakes in England and Wales have small catchments not directly linked to main river systems into which most STW discharge.
Three broad groups of sources can be identified, consented discharges (mostly STWs), non-point (diffuse) sources and a third group of small effectively point sources such as overflows from septic tanks, run-off from farmyards, road/tracks and other hard surfaces (Edwards and Withers 2008).
Where larger STWs are involved the introduction of tertiary treatment to remove phosphorus can result in a substantial reduction in phosphorus, which can lead eventually to the restoration of shallow lakes (Dudley et al. 2011; Phillips et al. 1999). In shallow lakes recovery often takes in excess of fifteen years, as a result of internal sources of phosphorus in the sediment (Carvalho et al. 1995; Jeppesen et al. 2005b). However, the introduction of phosphorus removal to significant discharges in a lake’s catchment remain one certain method of reducing phosphorus load significantly (Phillips et al. 2005).
Reducing diffuse sources of phosphorus is considerably more challenging. Small point sources are likely to deliver higher concentration of phosphorus (Edwards and Withers 2007) and give rise to phosphorus concentrations in feeder streams significantly in excess of 100 µgl-1 (Jarvie et al. 2008; Withers et al. 2009). Depending on the size of the lake these small point sources may be significant and it is appropriate to identify them. Single spot samples from streams within the catchment should be treated with caution, but synoptic surveys which sample several points in a small catchment can often reveal areas of elevated phosphorus concentration, (for example see Biggs et al. 2014) and these “hot spots” should be investigated to identify potential sources.
Septic tanks and small private treatment systems (PTS)
In rural areas overflows from septic tanks can be a significant source of phosphorus with the highest concentrations found in periods of low flow (Withers et al. 2011). In addition to overflows, phosphorus retention in the soakaways of septic tanks can be small, and is likely to reduce with time as soil adsorption sites become saturated (Beal et al. 2005), thus settlements not served by main sewers are potential significant sources of phosphorus (Johnes et al. 1996). It is difficult to accurately assess the potential impact from septic tanks, but estimates of population combined with typical loss coefficients can be used to assess the potential overall load and thus relative contribution to a lake (Brownlie et al. 2014; May et al. 2011; Withers et al. 2012).
Septic tanks retain up to 50% of input phosphorus through settlement in primary treatment and there is potential for further removal in the drainage field of soakaways (May et al. 2011). In contrast small treatment plants, not designed to remove phosphorus, often discharge direct to streams and thus have effectively higher phosphorus loads (May et al. 2011). Therefore, replacing septic tanks by small private treatment plants (PTS) with a direct discharge is not always the best solution. Planning policy can encourage the use of PTS that include tertiary treatment to chemically precipitate phosphorus, by linking rural development to the need to improve 3rd party treatment systems (Brownlie et al. 2014). However, investigations of a small number of different PTS showed no significant difference in the phosphorus concentration of the effluent from those with only primary (septic tanks), secondary or tertiary treatment (Brownlie et al. 2014). Thus there are no simple solutions to phosphorus arising from domestic rural sources. Awareness raising about the importance of good management of PTS are reported to reduce phosphorus concentrations in receiving waters (Bergfur et al. 2012). In some cases the most appropriate long-term approach might be to seek resources to provide a main sewerage system with treatment at a larger works where effective phosphorus removal can be achieved. Where this is not possible the siting of PTS as far away as practical from streams and particularly lakes is likely to be the best solution.
GP Add something about reed beds?
Reducing agricultural phosphorus loading sufficiently to reduce lake phosphorus concentration is challenging. There is a substantial literature investigating pathways of phosphorus loss and many models for use at different scales have been produced. The majority report a relatively wide range of estimated agricultural load which are spatially and temporally very variable, reflecting different sources and the many transport pathways (Hodgkinson and Withers 2007). They show however that animals and their waste products such as manure are the most important agricultural nutrient sources with source phosphorus concentrations much higher than that from drainage of agricultural land (Edwards and Withers 2007). The effective loading from animals, like that from human effluent i.e. PTS (see above) depends on connectivity. Thus direct run-off from farmyards, direct input of pig slurry for example, will be very substantial phosphorus sources. With good agricultural practice direct run-off should not occur and diffuse agricultural phosphorus control tends to be focussed on source minimisation and control of transport pathways.
Phosphorus transport tends to be associated with high rainfall events, from fields with low crop cover such as maize during winter (Withers and Bailey 2003) and where soil structure is compromised (Perks et al. 2015). Thus minimising phosphorus load is associated with improving soil structure, reducing soil erosion and slowing up hydrological pathways. Source control, with careful nutrient management planning at the field scale with allowance made for nutrients in animal manures, is also important to minimise the availability of phosphorus (Sharpley et al. 2015). There is an increasing emphasis on the identification of critical source areas to target edge of field remediation such as buffer strips, wetland and settlement basins, all designed to reduce run-off velocity and thus particulate transport’
Mitigation measures for agriculture
A list of potential mitigation measures to control diffuse nutrients applicable to England and Wales has been developed (Newell-Price et al. 2011) and combined with a spreadsheet tool “FARMSCOPER” which estimates baseline pollutant losses and can assess the impacts of different combinations of measures (Gooday et al. 2014). The mitigation measures target all three key aspects of diffuse pollution, source, mobilisation and delivery, and the tool uses well established national nutrient models to develop a library of export coefficients expressing nutrient loss for representative farm systems that reflect differing climatic and soil types in England and Wales. The tool is used to produce a suite of potential control options using a genetic algorithm to select from the many possible options. An application of the tool to the Hampshire Avon demonstration test catchment, UK is described by (Zhang et al. 2012).
The challenge of reducing phosphorus
In the UK there are examples of shallow lake improvements that have occurred following substantial reduction of point sources (Dudley et al. 2011; Phillips et al. 2005) but in many cases change has been slow (Jeppesen et al. 2005b). In lakes, retention of phosphorus within lake sediments is a key issue and there is growing recognition that phosphorus retention in the landscape is also an important consideration and that we are dealing with a landscape where phosphorus has accumulated for many years (Sharpley et al. 2009). Under certain conditions P remobilises and as contemporary concentrations of phosphorus are reduced these phosphorus sinks become sources releasing historic phosphorus. This is widely recognised in relation to lake sediments, but it also applies at the landscape scale (Sharpley et al. 2013). Thus reducing phosphorus must be seen as a long-term measure, it is essential if a currently eutrophic lake is to be restored. Point sources offer potentially easy wins, but reducing the wider background loading requires long-term changes to land-use management.
Isolation from catchment
For some smaller lakes isolation of the lake by diversion of inflow streams may appear to be a simple solution to reducing nutrient load. However, it is important to ensure that there is sufficient water available to continue to flush the lake. Where there are several inflows, diversion of one around or away from the lake may be beneficial, but if the main inflow is diverted conditions in the lake may deteriorate due to lack of flushing. A good example of an unsuccessful attempt to reduce nutrients occurred at Alderfen Broad, Norfolk UK. Phillips (1977) constructed a nutrient budget for this small very shallow lake and demonstrated that the main source of phosphorus was through a small inflow stream. The source of this phosphorus was thought to be from a small rural population and as a result the inflow stream was diverted around the lake. However, the result was that phosphorus concentration in the lake increased substantially as a result of release of phosphorus from the sediment during the summer when there was in effect no water exchange (Moss et al. 1986; Perrow et al. 1994). A more successful example was from another small very shallow lake, Upton Broad, Norfolk UK. Like Alderfen Broad this lake is isolated from the main river system, but its catchment was largely unimproved grazing or wet woodland and no significant rural population. As a consequence it had very low phosphorus concentrations (Bennion et al. 2001). In the mid-1980s, the unimproved pasture was converted to arable, but as a precaution drainage from this land was intercepted by a new ditch system that diverted drainage around the lake. Since then the lake has remained largely unchanged and maintained its low phosphorus and phytoplankton biomass (Phillips et al. 2015). These two contrasting examples highlight the need to consider the history of the lake when selecting management options. Alderfen Broad had been receiving elevated levels of phosphorus for many years and thus had built up a reservoir of phosphorus in its sediment which was released during the summer (Osborne and Phillips 1978). Reducing the phosphorus load by diversion of the inflow in this case exacerbated the situation as the historic nutrient load continued to express itself as a high phosphorus concentration during the summer and the lack of water exchange prevented this phosphorus from being lost from the lake basin. In contrast Upton Broad did not have a significantly elevated phosphorus load and thus the diversion to prevent an increased load was successful.
The recovery of shallow eutrophic lakes is nearly always delayed by the release of phosphorus from the sediment. The surface sediment needs to re-equilibrate and for a period following the reduction of external loading the rate of release of phosphorus can increase. In well flushed lakes this results in a net export of phosphorus from the lake and with time the phosphorus pool in the sediment is reduced and the lake is able to achieve a new equilibrium between water and sediment (Phillips et al. 2005; Sharpley et al. 2013). The time taken for this depends on the conditions in the lake but is typically 10-15 years (Jeppesen et al. 2005b) and sediment removal is a technique that has been used to speed up recovery (Kleeberg and Kohl 1999).
Typically surface sediment is removed by dredging. however success in reducing lake phosphorus has been variable (Kleeberg and Kohl 1999), and there are few detailed long-term studies available for assessment. To be successful the newly exposed sediment should have a high phosphorus binding capacity (Søndergaard et al. 2007) and external loads should have been reduced. In the Broads National Park, 23 very shallow lakes have been subject to sediment removal. Where detailed monitoring of the sediment was undertaken it was clear that although the sediment P concentration decreased after dredging the effect was short lived with pre-dredging concentrations returning within 3 years and thus there was little evidence that in these lakes dredging decreased the potential for internal phosphorus loading (Phillips et al. 2015). Based on the current evidence sediment removal to reduce internal phosphorus loading appears to have limited success.
In many shallow eutrophic lakes submerged macrophytes may have disappeared and one of the objectives of restoration is likely to be their re-establishment. In this context sediment removal may have benefits by exposing propagules of macrophytes buried in deeper sediments. There is some evidence of this from a number of small lakes in the Broads, although not all lakes where sediment removal has occurred have shown a significant increase in macrophyte abundance (Phillips et al. 2015).
Chemical dosing is an alternative strategy to reducing phosphorus loss from sediments. Iron and aluminium compounds have been used, either by addition to the water or direct injection to sediments. The use of iron is relatively ineffective in the longer term as in many sediments decomposition is sufficient to cause reducing conditions which allows phosphorus to be released (Foy 1985; Foy and Fitzsimons 1987). The use of aluminium and calcium is more likely to be effective and Mehner et al. (2008) reported that following dosing with these chemicals the eutrophic Lake Tiefwarensee recovered. The lake had been subject to external nutrient load reduction for a period of 10 years, but it required chemical dosing to reduce the internal nutrient load. However the reported monitoring of this lake only continued for a single year after dosing and thus the long-term influence cannot be assessed.
Recently there has been growing interest in the use of lanthanum modified bentonite clay
Add from (Spears et al. 2013) and probably (Lurling and Faassen 2012)
Conclude with (Spears et al. 2014)
Potential long-term disadvantages, locking in past pollution.
(Jeppesen et al. 2007) overview paper
Monitoring the response
Shallow lake restoration case studies
Key Restoration Messages
- Nutrient load reduction is the key to restoring nutrient-enriched shallow lakes. Look to remove/reduce inputs from upstream sewage treatment works and farmland as this is the only long-term, sustainable solution
- Do not undertake sediment removal unless nutrient-enrichment problems have been sufficiently tackled. Often the increase in water depth that results from sediment removal can make macrophyte recovery even less likely.
- Biomanipulation is only a sustainable approach to shallow lake restoration where nutrient-enrichment issues have already improved water quality status
- Avoid over-stocking of bottom-feeding angling fish such as common carp and bream. Do not use restored lakes for intensive wildfowl stocking. Both of these lake uses lead to turbid, conservation-poor lakes.
Shallow Lake Reference List
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