香京julia种子在线播放

Two datasets were identified as containing records of Zostera from multiple sites. The OSPAR Commission (2017) dataset represents the current known areal extent of seagrasses in the United Kingdom and includes records on Zostera between 1986 and 2015. Under the WFD, a range of government agencies (e.g., English & Northern Ireland Environment Agency, Scottish Environment Protection Agency, and Natural Resources Wales) are required to assess the condition of seagrass to help determine the biological condition of United Kingdom water bodies (Foden and Brazier, 2007). The outcome of this is another dataset that includes areal extent of Zostera meadows monitored under the WFD between 2007 and 2017. These data were analysed by region and date on QGIS (version 3.2.1) and were shown to contain substantial gaps. To supplement these, we contacted stakeholders from a multitude of organisations targeting local councils, national and devolved governments, government advisory organisations, fisheries authorities, private environmental consultants, and scientists who work on seagrasses in the United Kingdom. From these searches, 14 additional contributors supplemented the OSPAR and WFD datasets, the collective of which makes up all the known available data, based on the searches we undertook (see Supporting information). Since species identification was not provided across all data sets, they have not been included here. However, as an intertidal species, there are far less technological constraints associated with surveying Z. noltii. Because of this we expect it to be in the majority, and further expect many Z. marina meadows to have gone unreported. Because of these inconsistencies, we have decided not to discriminate between species, since greater abundance of one species is likely due to mapping inconsistencies rather than variances in the conditions that allow one species to proliferate over another. It should also be noted that Zostera angustifolia was once considered its own species, although is now recognised as a phenotype of Z. marina (Becheler et al., 2010). As we do not distinguish between species, Z. angustifolia is treated in the same way as Z. marina and Z. noltii.

Data were provided in the form of individual observations (point data) and area estimates (polygon data). Polygon data were used to provide the area estimates contained herein. Spatial assessments were made using QGIS (version 3.2.1) and all data were analyzed for duplicates or overlaps. Where they occurred the most recent data was used, unless differences between years occurred that represented data collection restraints rather than area changes. For example, when data were present in the same location from 2016 to 2018 but the 2018 data held substantially reduced area, it was assumed that this was not representative of habitat degradation but restrictions to accessing the full extent of the meadow. This was supported by several data fluctuating between 3 years – e.g., 2014 and 2016 showed the same area cover, but 2015 showed far reduced cover. The area of each polygon was calculated in m² using a cylindrical WGS-84 projection and converted to hectares (ha) for reporting purposes.

The contemporary data represent the minimum area of seagrasses in the United Kingdom since some meadows have certainly gone unreported. OSPAR data were used to provide high and medium certainty estimates of historic mapped areal extent. The maximum seagrass extent for each record within the dataset was checked against contemporary records and where contemporary records were found the difference between largest (oldest) and current meadow size was used to provide high certainty loss estimates. Where no contemporary record of the meadow was found, these were considered as spatial loss and included in medium certainty loss estimates. We acknowledge our approach, and the subsequent estimates of changes in coverage through time, is constrained by sampling efforts and data reporting of past research efforts, but this represents the best use of the best available data.

Systematic reviews are used to encapsulate a broad range of literature on a discrete subject by aggregating large data and rigorously extracting relevant information (Minx et al., 2017). We followed the distinct protocols required to achieve a systematic review by: (1) defining the discrete subject parameters and the timeframe of interest; (2) creating a search term to encapsulate all data that might be relevant to the subject; (3) inputting this into Web of Science (Thompson Reuters) to extract a literature database; (4) justifying and making a transparent selection of the literature; and (5) providing a synthesis of the relevant literature. Full details of these methods are provided in the supporting information (SI).

Web of Science includes published, peer-reviewed articles as far back as 1990. Because of the distinct lack of published data on seagrass area cover in the United Kingdom, and a need to capture data as far back in time as possible, it was necessary to broaden the search to include published, unpublished, and grey literature. These data were collected by extensive internet searches, through contacting stakeholders from government, private organisations, and NGO’s, and from scientists and the public who work in seagrass science, conservation, and management throughout the United Kingdom. Papers were qualified and data extracted based on the same criteria as the systematic search (SI). Both search unearthed 179 papers that were considered relevant to this work.

Because of the scarcity of historic empirical data, and the observation that many of the early qualitative reports are almost certainly subject to SBS, we used available data to model the maximum historic extent and low certainty loss estimates of seagrasses in mainland Britain.

In 1991, the Nature Conservancy Council (NCC) undertook a report on the 155 estuaries that exist in mainland Britain (Davidson et al., 1991). In 1932, Butcher reported on the distribution of Zostera in the United Kingdom, including a spatial distribution map of mainland Britain that corresponds very closely with the estuaries map presented in the NCC report (Figure 1; Butcher, 1934). Further, qualitative data suggest that before WWI seagrass would once have been found across a large proportion of subtidal mud- and sandflats and on the lower ranges of most intertidal flats throughout the United Kingdom, especially prevalent in estuaries (Cotton, 1933; Butcher, 1934, 1941; Cottam, 1934). Sub and intertidal mud- and sand-flats, in particular, estuarine ones are, therefore, a good proxy for modelling historic seagrass distribution.

We identified locations with best available data on seagrass area cover from either historic or contemporary estimates, where contemporary estimates represent meadows that are in reasonably good condition, and where total mud- /sandflat area for each site was available. This restricted the inclusion of sites to those designated as Special Protected Areas (SPA’s) and Special Areas of Conservation (SAC’s; JNCC, 2005, 2015, 2018) since such sites have been accurately mapped by the Joint Nature Conservation Committee (JNCC; Table 1).

The chosen sites represent typical environmental variation for United Kingdom seagrasses including intertidal and subtidal habitats within estuaries, rivers, lochs, and spits. Seagrass area coverage per hectare of mud- / sandflat was calculated for each site by dividing total seagrass area by total mud- /sandflat area (Table 1; Figure 2i). Seagrass coverage was estimated using bootstrapping techniques (Manly, 2006). Our 10 sites were randomly selected with replacement 1,000 times and multiplied by total mud-/sandflat area estimates from (a) the OSPAR dataset (2017), which includes data in and outside of estuaries, and (b) the NCC report (Davidson et al., 1991), which includes data from estuaries only (Figures 2a,b). The average of these 1,000 estimates were used to estimate maximum extent of seagrasses around mainland Britain (Figures 2ai,bi). We used this simple bootstrapping procedure (Manly, 2006), rather than more typical parametric statistical methods, due to the paucity of available data. The data provide low certainty maximum seagrass extent around mainland Britain. In addition to the simple modelling approach, datasets were also obtained from previous studies using habitat suitability modelling to estimate potential seagrass distribution (Brown, 2015; Defra, 2020).

The total mapped areal extent of contemporary seagrass records (post-1997) from the OSPAR dataset, the WFD dataset, and all other contributors includes 47 surveys spanning 20 years, 79% of which are from the last 10 years (see Supporting Information). In total, the data confirm the presence of 8,493 ha of seagrasses in the United Kingdom (Table 2). Occurrence of seagrasses is not uniform. Half of all mapped seagrass occurs in the Scottish Highlands (20%), Devon (16.2%), and Northern Ireland (14.3%; Table 2). Seagrass occurrence ranges from patches less than 1m² to meadows up to 1,200 ha (i.e., Cromarty Firth, East Scotland). The average size of seagrass record is 2.64 ± 32.22 ha.

The OSPAR dataset, which represents the currently used known areal extent of seagrasses in the United Kingdom, includes 13,753 ha of seagrass. Of this, 8,835 ha (64.2%) was recorded pre-1998 and 4,919 ha (35.76%) was contemporary (post-1997). Within the OSPAR dataset, there is an inverse relationship between average meadow size relative to age of record, i.e., the average area of historic seagrass meadows is 71 ± 218 ha whilst the average area of contemporary seagrass meadows is 2 ± 238 ha. The average for all the meadows included in the OSPAR dataset between 1986 and 2015 is 4 ± 50 ha. GIS point data exist beyond the extent of these mapped areas but the extent of any seagrasses associated to such an observation is unknown and therefore not included within this analysis.

The total mapped historic extent of seagrasses in the United Kingdom is 16,524 ha. The total documented loss of seagrasses since 1936 is 6,697 ha. A further 1,364 ha of seagrass habitat has not been revisited since 1998 (Table 3), a disproportionate amount of which is from Scotland (Table 4). With high certainty the UK has, therefore, lost 44% of its seagrass since 1936, 34% since the 1980s. With medium certainty, including historic data with no recent observations, 50% has been lost since 1936, and 42% since the 1980s.

The first published account of seagrass in the United Kingdom that we are currently aware of was in 1831 (Winch, 1831), where it was included in a publication on the “Flora of Northumberland and Durham.” This observation is from a location on the Tyne River long since reclaimed and now an industrial estate.¹ A peak in publications occurred around the time of the 1930’s wasting disease when naturalists became concerned with the substantial degradation of sites throughout the United Kingdom (Figure 3). Publications were sporadic until 1990 (n = 20) and since then have occurred more frequently as a series of peaks and troughs (n = 66).

Attempts to describe seagrasses across the United Kingdom were made by Butcher in 1932, but without any defined methodology (Butcher, 1934). He considered the occurrence of seagrasses regionally but did not provide area estimates. Seagrasses seemingly occurred ubiquitously, “apart from wave-swept, shingly and rocky shores to the west of the country” (Butcher, 1934). The abundance of seagrasses in sheltered and protected areas on the east coast was noted, as were plentiful populations in the lochs of Ireland and the west of Scotland (Butcher, 1934). Aside from this early attempt, efforts were made to document the status of seagrasses in Scotland in 1933 (Cleator, 1993), in Wales (Kay, 1998), and in Cornwall and the Isles of Scilly in 2002 (Hocking and Tompsett, 2002a,b). These reports provide presence and absence data and indicate widespread declines, but do not provide usable spatial estimates of area.

A full analysis and synthesis of the systematic search is provided in Supplementary Information. The literature analysed through the systematic review show ubiquitous declines across almost every region of the United Kingdom. Areas where good historic data are available show declines of between 40% (Cornwall) and 100% (Suffolk; Table 3). Although historic quantitative data are rare, the ubiquitous declines in seagrass areal extent are evident. Further, these historic declines are matched by pervasive recent declines, which suggest we are yet to stem this trend.

The proportion of seagrass area per hectare of sub- and intertidal mud- /sandflat ranged from 4 to 86% with an average of 32 ± 27%. The 1991 NCC report (Davidson et al., 1991) established that estuaries comprised a total of 530,000 ha of coastal waters in mainland Britain. Of these, half are in England and almost one third are found within Scottish waters (Davidson et al., 1991). Within these, mud- /sandflats make up about 43%, totaling 254,400 ha (Davidson et al., 1991). The OSPAR dataset does not include any data from Ireland and is lacking in Scottish datapoints. It reports 143,571 ha of sub- and intertidal mud- /sandflats in mainland Britain, including those outside of estuaries. Considering one third of estuaries is found in Scotland it is unsurprising that these figures do not align. The total current mapped areal extent of seagrasses in mainland Britain (from this papers data) is 6,760 ha.

Using the NCC data on total mud- /sandflats area, the estimated maximum seagrass extent for mainland Britain is 81,953 ha, with an upper 95% CI ranging from 126,430 to 40,964 ha (Table 5; Figure 4). Using the OSPAR data on total mud- /sandflats area, the maximum seagrass extent for mainland Britain is 43,559 ha, with 95% CI ranging between 72,647 and 24,267 ha (Table 5; Figure 4). Statistical comparisons between the two seagrass coverage values were made by comparing 95% CIs. Overlapping CIs indicated no significant difference between area estimates. The modelled data suggests that, with low certainty, between 36,799 and 75,193 ha of seagrasses has been lost from mainland Britain, this would represent an 84 and 92% decline, respectively (Table 3).

Rationalising the probable causes of such vast losses of seagrass in Britain is at best difficult, mostly because robust estimates regarding historic spatial extent of Zostera are limited. Typically, the seagrass wasting disease Labyrinthula has been described as the primary cause of virtually all seagrass loss in the United Kingdom (Butcher, 1934; Cottam, 1934; Den Hartog, 1983; Garrard and Beaumont, 2014). We argue that this assumption is itself a result of SBS. This assumption stems from the discussion within Butcher’s report, where undoubtedly seagrasses were lost due to disease (Butcher, 1934). Butcher reported changes to habitat based on his own intertidal experiences of seagrass abundance, supported by those of individuals whose baselines only go as far back as their own inherited knowledge (Butcher, 1934; Pauly, 1995). This case of SBS means the basis on which Butcher referred to healthy populations of seagrasses is likely a gross underrepresentation of what once occurred in these waters and is an excellent example of how SBS impacts contemporary environmental knowledge (Butcher, 1934). There has been no enquiry as to whether the seagrass habitats Butcher was assessing were already heavily degraded, with almost all the literature pointing to this period for the cause of seagrass degradation (Butcher, 1934). Considering the early industrialisation of the United Kingdom and a long history of mining activity, it is almost certain that the systems Butcher assessed in 1930’s would have already undergone dramatic declines. His assumptions were also largely based on visits to sites on the eastern shores of England. It is likely that persistent gradual declines had been occurring for centuries before Butchers report and these have continued to the present day (Jackson et al., 2016; Jones and Unsworth, 2016; Unsworth et al., 2017; Jones et al., 2018).

As the first country to industrialise in the 17 and 18th centuries, Britain had been undergoing dramatic land-use transformation long before Butcher assessed the status of seagrasses (Butcher, 1934). Industrialisation is intrinsically linked to environmental degradation. By the time, Butcher had been writing, dramatic physical alterations to the United Kingdom landscape had been occurring for at least 300 years (Butcher, 1934, 1941). A reference to seagrasses in the Tyne estuary in the early 1800’s (Winch, 1831) referred to a site that has since been reclaimed, now containing an industrial estate. Coastal reclamation, dredging and building of sea walls were prevalent in the 17th century, 200 years prior to this account (Batty, 1997), and are highly likely to have been as in conflict with seagrass as they still are today (Erftemeijer and Lewis, 2006). Importantly, the UK was at the forefront of the global metal industry, with metal mining prevalent throughout many parts of the United Kingdom during the 1700s and 1800s, with many of these mines (e.g., Wales) still producing extensive metal contaminated discharge into coastal and estuarine waters today. The negative impacts of a suite of heavy metals, causing toxic conditions for seagrasses, are well documented (Prange and Dennison, 2000; Macinnis-Ng and Ralph, 2002). Many areas thought to be viable seagrass sites within Welsh habitat suitability models are areas of historic heavy metal mining contamination, such as NE Anglesey (Whiteley and Pearce, 2003).

In addition to industrial development, the scale of overfishing of oysters around the United Kingdom cannot be ignored as a fundamental change and major disturbance to the environmental conditions. There is growing evidence of the close positive interactions that occur between seagrasses and many bivalves (Perkins, 1988; Gagnon et al., 2020). Locations such as the Firth of Forth have entirely lost up to 5,000 ha of oyster beds (Thurstan et al., 2013). These oyster beds would have fundamentally influenced the volume of suspended particles in the water column and hence the water clarity, creating conditions suitable for photosynthetic production by seagrasses. Similar estimates are available for areas, such as the Solent, the Thames, The Clyde, the Humber, and the Severn (Thurstan et al., 2013). The extraction activities of these Oyster fisheries were also likely to have disturbed and remobilised sediments at a high frequency over major areas for prolonged periods, potentially negatively impacting seagrasses. It is not only Oyster fisheries that would have likely had a historic negative impact upon seagrass. As early as the 1700s, bottom trawling was already widespread around the coastal waters of the Great Britain (Jones, 2018) and such activity is well known to directly damage seagrasses (Blaber et al., 2000). Seagrasses remain threatened in the United Kingdom today by activities which reduce water quality, and direct physical disturbance (Jones and Unsworth, 2016; Green et al., 2018).

The present analysis highlights a lack of coherent and systematic monitoring and mapping programmes of seagrass meadows in the United Kingdom. That 64% of records in the OSPAR dataset are older than 20 years highlights the prolific lack of constant effort in seagrass mapping. Despite this fact, the OSPAR dataset is the baseline for estimates of United Kingdom seagrass extent included by the UNEP World Conservation Monitoring Centre (WCMC). Since this is the first attempt to provide an accurate map and up-to-date map of seagrass occurrence and declines for an entire country it is likely that these are not the only data included by WCMC that are inaccurate (McKenzie et al., 2020). The paucity of current data means that estimates, even coarse ones, are a necessary step to evaluating the pressures imposed on seagrasses, in keeping with the accepted approach of “use available data” (Hiscock, 1997). However, data inconsistencies can make it challenging to talk meaningfully about global seagrass trends and arguably managers should only be working with temporal and spatial data that we are reasonably confident is accurate. Regional and local mapping of sites around the world is important in ensuring that current attempts to increase seagrass abundance through restoration, rehabilitation, and conservation have any hope to succeed.

Data inconsistencies found within this work are most obvious in Scotland, where most data were collected over 20 years ago. Although the isolation from human population likely means there is potential for these meadows to remain intact, many have likely been impacted by the extensive and continually expanding salmon aquaculture industry, with fish farming a known cause of seagrass loss in other parts of the world (Berry and Davison, 2001). The meadows where historic and contemporary data are available show mass declines. The once huge meadows in the Cromarty and Dornoch Firths have been reduced from 3,241 to 1,200 ha and 2,546 to 117 ha, respectively. Regardless, huge swathes of Scottish waters have not been surveyed for over 2 decades and could represent a vital stronghold of this once ubiquitous United Kingdom habitat. Their condition and extent should be assessed with urgency. Scotland is not the only region where survey efforts since 1998 have been insubstantial. Pre- and post-1998 maps show a reduced survey effort across all regions (Figure 5).

Recent efforts have been made to estimate the amount of carbon stored within the United Kingdom seagrass meadows (Röhr et al., 2018; Green et al., 2018; Lima et al., 2020). These papers analysed sediment from 13 meadows along the southwest coast of the United Kingdom for organic carbon content. For areas around the Solent and adjacent harbours, Lima et al. (2020) found 33.8 ± 18.5 Mg C ha⁻¹ in the top 30 cm of sediment. Whilst Green et al. (2018) found meadows contained a total of 141 ± 73 Mg C ha⁻¹, within the top 1 m of seagrass sediment. We have used mean values across both United Kingdom seagrass species as current carbon values are limited and explicit assessment of the impact of species and environmental variation upon this are largely absent. Based on these average figures, the estimated total carbon stored in the top 100 cm of recently mapped seagrasses of the United Kingdom is ~1.2 Mt carbon. In mainland Britain, this figure is 0.9 Mt of carbon (Table 5). Based on the upper (81,952 ha) and lower (40,965 ha) estimates of historic seagrass distribution, mainland United Kingdom seagrass meadows could once have contained between 5.7 and 11.5 Mt of carbon (Table 5). The upper value of this is equivalent to 3% of the United Kingdoms CO₂ emissions in 2017 (Eaton, 2019).

There exists only one value currently of United Kingdom seagrass carbon sequestration, but this value itself is an estimate (Garrard and Beaumont, 2014), hence there are no reliable estimates of seagrass sequestration rates in the United Kingdom. Reasonable and frequently used rates in the literature give low (0.044 cm yr.⁻¹), medium (0.202 cm yr.⁻¹), and high (0.42 cm yr.⁻¹) bounds to frame carbon sequestration estimates (Duarte et al., 2013; Lavery et al., 2013; Macreadie et al., 2013; Miyajima et al., 2015; Röhr et al., 2018). Here, sequestration rates were estimated by dividing the total carbon estimates by the amount of time it takes to accumulate this stock using the sedimentation rates above, to provide estimates on average annual carbon accumulation of United Kingdom seagrass meadows (Lavery et al., 2013; Röhr et al., 2018; Table 6.). Assuming a medium sedimentation rate, the seagrass meadows of the United Kingdom are accumulating roughly 0.024 Mt C yr⁻¹ (Table 6). Assuming this medium sedimentation rate, historic undisturbed seagrass meadows of the United Kingdom could have been absorbing 0.232 Mt C yr⁻¹ (Table 6).

Considering the need to include natural ecosystems in climate mitigation strategies, there is increasing interest in placing monetary valuations on carbon stock and sequestration estimates. The United Kingdom government has recently implemented a legal commitment to achieve Net-Zero greenhouse gas emissions by the year 2050. To reach this target will require major economic reforms and substantial increases in natural carbon sequestration capacity. The current United Kingdom carbon value is £24/t C (DECC, 2011; Green et al., 2018). However, according to the Grantham Institute, a price that is consistent with the Net-Zero targets needs to begin at £50/t C, rise to £75/t C in 2030, and to £160/t C in 2050 (Burke et al., 2019). Based on today’s market price, the value of the carbon stored in the top 100 cm of recently mapped seagrass stands at a value of £29 million with yearly (medium) sequestration value (Table 6) of £0.58 million, rising to £3.9 million over the next 30 years, if the projections hold true (Table 7). Taking the upper (Table 6) range of the historic estimates of seagrasses in mainland Britain, at today’s market value, these would once have contained £276 million worth of carbon in their sediments. In this undisturbed state, these seagrass meadows could have been responsible for sequestering £5.6 million worth of carbon every year, rising, if the projections hold true, to £37.2 million over the next 30 years (Table 7). These figures, although crude, offer a powerful indicative snapshot of what has been lost through long-term environmental degradation, and support the need to offer protection to those seagrass meadows that remain. Seagrasses use to be ubiquitous along the shores of the United Kingdom, and if restored to even part its former extent; this ecosystem will provide valuable support to reaching a carbon neutral future.

The value that seagrasses provide for other ecosystem services cannot be readily quantified financially with current data; however, the functions they play are extensive in the United Kingdom, particularly with respect to fisheries support, biodiversity, nutrient cycling, and sediment stabilisation (Nordlund et al., 2016). United Kingdom studies have revealed that seagrass harbours 4.6 times the abundance of fish of unvegetated habitat at a density of 6,000 fish per hectare (Bertelli and Unsworth, 2014), resulting in United Kingdom seagrasses currently supporting approximately 50 million fish, many of commercial importance (e.g., juvenile whiting, cod, plaice, and pollack). Based on the upper (81,952 ha) and lower (40,965 ha) estimates of historic seagrass distribution, mainland United Kingdom seagrass meadows could once have contained between 200 and 400 million fish, a potential loss of 274 million fish based on the comparison to unvegetated seabed.

Given the eutrophication issues faced by many coastal waters around the United Kingdom, the loss of seagrasses will have had a detrimental impact upon water quality. For example, Z. marina meadows may cycle approximately 49 kg of nitrogen per hectare year (Watson et al., 2020). This equates to the current annual cycling of 416 tonnes per year that could have historically been as high as 4,015 tonnes per year. Seagrasses also help reduce coastal erosion through improved stability of sediments. The extent of this role is often context dependent (Ondiviela et al., 2014) but available evidence indicates that at high density Z. marina may increase by up to 10 fold the sheer strength of the sediment (Widdows et al., 2008). Given the vast loss of seagrasses in eastern parts of the United Kingdom in areas know to be subject to coastal erosion and predicted to suffer from increasing impacts of rising sea levels, we hypothesise that at its historic extent seagrasses would have played a pivotal role in reducing coastal erosion.

The authors declare that the research was conducted in the absence of any commercial or financial relationships that could be construed as a potential conflict of interest.