A stock–recruitment model with a temperature component was used to estimate the effect of an increase in temperature predicted by climate change projections on population persistence and distribution of twaite shad Alosa fallax. An increase of 1 and 2° C above the current mean summer (June to August) water temperature of 17·8° C was estimated to result in a three and six-fold increase in the population, respectively. Climate change is also predicted to result in an earlier commencement to their spawning migration into fresh water. The model was expanded to investigate the effect of any additional mortality that might arise from a tidal power barrage across the Severn Estuary. Turbine mortality was separated into two components: (1) juvenile (pre-maturation) on their out migration during their first year and on their first return to the river to spawn and (2) post-maturation mortality on adults on the repeat spawning component of the population. Under current conditions, decreasing pre-maturation and post-maturation survival by 8% is estimated to result in the stock becoming extinct. It is estimated that an increase in mean summer water temperature of 1° C would mean that survival pre and post-maturation would need to be reduced by c. 10% before the stock becomes extinct. Therefore, climate change is likely to be beneficial to populations of A. fallax within U.K. rivers, increasing survival and thus, population persistence.
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How species respond to climate change is largely unknown. Changes in distribution, phenology and abundance have been described (Hughes, 2000; Parmesan, 2006) and can have positive or negative consequences for the persistence of a species. Not surprisingly, commercially and economically important species and those species that are significantly below their reproductive potential have been of primary focus when attempting to understand responses to climate change.
As part of the United Nations Framework Convention on Climate Change (UNFCCC, 1992) and the Kyoto Protocol (UNFCCC, 1998), industrial nations agreed to reduce their greenhouse gas emissions by an average of 5% (compared to 1990 levels) by 2012 (Liverman, 2008). To achieve this, a reduction in energy consumption coupled with an increase in energy production from renewable sources is necessary (Mitchell, 2006). Renewable energy sources include wind, sunlight, geothermal heat, tides and rain. Such sources are considered as green energy sources as they are continually replenished (renewable) and, therefore, generally not considered to have an adverse effect on the environment. The processes required to harness some sources of renewable energy, however, can have deleterious effects on the surrounding environment, e.g. wind farms affecting birds (Exo et al., 2003) and hydropower schemes affecting fishes (Schilt, 2007).
In the U.K., tidal power has the potential to produce at least 10% of the country's electricity needs, of which 4·4% of the supply could come from a tidal power barrage across the Severn Estuary (SDC, 2007). Such a scheme would make a considerable contribution to the U.K. government's aspiration that 20% of the national electricity supply should come from renewable sources by 2020 (HMSO, 2007). The Severn Estuary, however, is subject to a number of national and international designations. It is classified as a Special Protection Area (SPA) for birds under the EU Birds Directive (E.U., 2009) and is a Special Area of Conservation (SAC) under the EC Habitats Directive (E.U., 1992) as are the Rivers Wye and Usk, which drain into the estuary. An initial assessment of the effects of a tidal barrage on the integrity of the SAC, a primary designation feature being twaite shad Alosa fallax fallax (Lacépède), indicated a risk of mortality sufficiently high to eradicate the stock (SDC, 2007). Alosa fallax has declined considerably in abundance throughout its geographic range (Baglinière & Elie, 2000; Aprahamian et al., 2003). For this reason, A. fallax has been listed under the international union for the conservation of nature (IUCN) World Red Data Book (IUCN, 2006), included in Appendix III of the Bern Convention (CE, 1979) and incorporated into Annexes II and V of the EC Habitats Directive (E.U., 1992).
Throughout Europe, there are several factors linked with the continued decline of A. fallax populations including: poor water quality (particularly in lower reaches and estuaries), habitat destruction as a result of flood defence and reengineering works or gravel extraction affecting available spawning habitat (Aprahamian et al., 2003). Perhaps the most significant effect, however, is considered to be due to artificial barriers. Structures such as dams prevent adults from accessing spawning grounds and their effect may be exacerbated by overfishing as fish congregate and become easier to capture below obstructions (Baglinière & Elie, 2000).
In addition to those pressures, the possible influence of climate change must also be considered. In the U.K., self-sustaining populations of A. fallax are currently confined to rivers draining into the Bristol Channel of the west coast: Rivers Severn, Wye, Usk and Tywi (Aprahamian & Aprahamian, 1990). These rivers represent the northern limit of the known spawning distribution of A. fallax (Aprahamian et al., 2003) in the U.K. Within these regions, the medium emissions scenario (UKCP09; Defra, 2010) predicts the central estimate of increase in mean summer air temperature to be 2·5° C by 2050, ranging between 1·0 and 4·6° C albeit with a wide range of uncertainty (Murphy et al., 2009).
Alosa fallax is an anadromous species, entering the Severn Estuary in April to start the freshwater phase of their spawning migration (Aprahamian, 1981). The timing of their movement into the estuary appears related to temperature; peak immigration occurring at temperatures ranging between 10·6 and 12·3° C (Aprahamian, 1988). A close correlation between migration and temperature serves to minimize egg and larval mortality as well as increasing the probability of adult postspawning survival (Leggett, 1985). Spawning occurs between May and July with June being the main month (Aprahamian, 1982). The juveniles are present in the estuary from July until they migrate seaward in the autumn (Claridge & Gardner, 1978; Aprahamian, 1988). A portion of the 1 year-old fish re-enter the estuary in the spring before again migrating seaward in the autumn (Aprahamian, 1988).
The introduction of tidal barrages to generate energy can increase mortality rates for species that pass through the turbines as part of their migration (Stokesbury & Dadswell, 1991; Gibson & Daborn, 1995; Gibson & Myers, 2002). In the context of global warming and its ecological consequences (Hughes, 2000; Parmesan, 2006), temperature is expected to have a positive effect on the persistence of A. fallax, but in the Severn Estuary any benefits that might arise from an increase in temperature (Holmes & Henderson, 1990; Aprahamian & Aprahamian, 2001) may be offset by an increase in mortality from the tidal barrage scheme. The aim of this study was two-fold: (1) to quantify the change in population abundance that might arise from an increase in water temperature in line with current climate change predictions, and (2) what degree of additional mortality, as a result of barrage introduction, would be responsible for a decreasing population persistence under varying climatic conditions.
Materials and methods
The adult population of A. fallax entering the Severn Estuary at the start of the freshwater phase of their spawning migration was sampled between 1979 and 1997. Counts of A. fallax were obtained from the catches of Atlantic salmon Salmo salar L. putcher net fishermen operating near Lydney on the Severn Estuary (51·689° N; 2·564° W) (Fig. 1) between 15 April and 15 August. The putcher rank consists of 650 conical shaped traps constructed from metal bar 6 mm in diameter, designed to sieve fish on both the flood and the ebb tides (Aprahamian, 1981). Between 1979 and 1988, the size of the traps and their arrangement in the rank changed; after 1988 fishing effort was constant. The changes affected the size of the opening of the traps as opposed to the spacing of the bars and thus, their selectivity was considered not to be affected. The effect of the changes on fishing effort and the raising factors used to compare catches taken between 1979 and 1987 with those after 1987 are shown in Table I. In 1982 no sampling was possible as the putcher rank collapsed and for the years 1983 and 1984, no estimates of fishing effort were available. Sampling was conducted on a daily basis to ensure both tides were sampled, to avoid bias from variation in diurnal patterns of behaviour or catchability (Aprahamian, 1981).
Table I. Change in the effective fishing effort of the putcher rank over the period 1979 to 1997, together with the raising factors used to adjust the catch to the effort deployed between 1988 and 1997
Effective fishing effort (arbitrary units)
Sampling was carried out between 15 April and 19 June, the main migration period through the estuary (Aprahamian, 1981). Samples of the catch or sub-samples (50 fish) of the catch if the number of fish caught exceeded 50 were taken in order to partition the run according to sex, age and spawning history. Age and spawning history were determined from analysis of scales (Baglinière et al., 2001). As the timing of the migration through the estuary differed between years (Aprahamian, 1988), the migration was divided up into eight periods (week) of either 13 or 14 tides and one of 18 tides. The latter was the first period and took account of the number of days required to install the putcher rank. The periods were ranked according to catch per unit effort (CPUE) and the top five periods used to calculate an index of the size of the spawning stock (mean catch per tide). This was because in some years no, or relatively few, samples were taken towards the end of the run. The available tides sampled ranged from 5 to 67% year−1(mean = 31% year−1) and the number of fish aged ranged from 110 to 617 fish year−1(mean 359 fish year−1).
The CPUE (X) for each age class (i) was calculated separately for males and females as follows:
where x = mean catch of A. fallax per tide of fish age i, spawning number j (spawning number 1 is a fish spawning for the first time or virgin spawner, spawning number 2 is a fish spawning for the second time it will have one spawning mark on its scale, etc.) in period k.
To reduce the possibility of bias derived from sampling too few tides, the index was estimated using the change in abundance between successive age groups. Subjectively a value of 10% of the 127 tides available between 15 April and 19 June was taken as the threshold. In those years where <10% of the tides were sampled (1979 and 1982 to 1984), the index was calculated as follows: Δij = lnXij(t)−lnXij(t+1), where Δij = instantaneous rate of change between successive age groups for fish of a particular age (i) and spawning number (j), Xij = the mean catch per tide for the five periods, of fish of a particular age (i) and spawning number (j) and t = years.
In 1979, CPUE was backcalculated using the 1980 CPUE data and the instantaneous rate of change between the years 1980 and 1981. For the period between 1982 and 1984, the CPUE for fish age 6 years and older in 1982, 7 years and older in 1983 and for fish aged 8 years and older from 1984 were estimated from the 1981 CPUE data and the instantaneous rate of change between the years 1980 and 1981. The other age classes were backcalculated from the 1985 CPUE data and the instantaneous rate of change between the years 1980 and 1981 and between 1985 and 1996. This was done to make allowance for any possible change in efficiency as a result of the change in construction material and because it provided the closest approximation to the age structure recorded in 1983 and 1984. Only samples where the number of fish from a particular age group and spawning number were ≥5 were used to estimate the instantaneous rate of change.
Selectivity (Ci) was estimated from the catch curve for each age group as follows: , where R = proportion of repeat spawners of age class i and S = instantaneous rate of spawning mortality (0·6651), derived from the catch curve.
The correction factors applied to each age were as follows: for age classes 3, 4 and 5 years a factor of 13·11, 3·64 and 1·81 was applied, respectively; for age classes 6 to 10 years a value of 1·00 was used in each case.
An index of the total number of eggs deposition (E) in year t was calculated as follows: , where the fecundity of age class i (Fi) is calculated as: Fi = 34324I0·538(Aprahamian,1982), where i = age in years.
The CPUE at age 6 years (X6) was used as a proxy for recruitment. Age at maturity can vary between cohorts by up to a year but, in general, >98% of fish have matured (Aprahamian & Lester, 2001). For the 1974 and 1975 year classes and those between 1979 and 1991, the age 6 years CPUE was determined directly from catch data using equation 1. For the 1973 and for the 1976 to 1978 year classes, the index was estimated from the CPUE and the instantaneous rate of change as outlined above. The 1972 index was estimated from the number of 7 year-old fish caught in 1979, assuming that 68% of fish age 6 years survive to age 7 years. The survival rate was determined from the relationship between the number of fish age 6 years caught in year t and the number of 7 year-old fish caught in year t + 1, for the 1973 to 1990 year classes (Aprahamian & Lester, 2001).
Quantitative monthly sampling of juvenile A. fallax has been carried out at Hinkley Point B nuclear power station (Fig. 1) since October 1980 (except during 1986) (P. Henderson, pers. comm.). Sampling dates were chosen to coincide with intermediate range tides in the spring–neap cycle. Sampling was standardized, so that on each visit six consecutive samples were collected over a 1 h period using plastic baskets covered with 6 mm mesh and positioned to collect all the debris washed from two of the four drum screens which filter the cooling water entering the power station. The debris was sorted and the number of A. fallax captured per hour recorded and the standard length (LS) of the fish measured. The method is selective towards juvenile fish with the majority of A. fallax caught being of age 0+ years (Holmes & Henderson, 1990).
For the 1992 to 1996 year classes, the recruitment of age 6 year-old A. fallax was estimated from the relationship between the number of juvenile A. fallax caught at Hinkley Point B nuclear power station between 1 June in year t and 31 May in year t + 1, and the CPUE index of 6 year-old female shad caught in the putcher fishery in year t + 6 (Aprahamian & Aprahamian, 2001).
Stock–recruitment was modelled using the Ricker relationship (Ricker, 1954), , where Et = the number of eggs deposited in year t (stock), E6(t+6) = the number of eggs deposited by fish age 6 years in year t + 6 (recruits), a = egg survival at low density and b = index of density dependence.
Water temperature is strongly correlated with spawning success, explaining 77% of the variability (Aprahamian & Aprahamian, 2001). Changes in temperature, therefore, have the potential to greatly affect the number of A. fallax returning to a river. Under existing (baseline) climate scenarios, interannual success can be highly variable, whereby temperatures can be sufficiently low to be detrimental to recruitment or alternatively, high enough to support population persistence (Aprahamian & Aprahamian, 2001). In order to assess the role of temperature in recruitment success and the potential for climate change to offset any increased rates of mortality associated with the introduction of tidal power turbines, recruitment success under baseline and climate scenarios were modelled. Uncertainty was inbuilt within the model by incorporating a stochastic temperature function (T; equation 2) based on daily water temperature data obtained from Oldbury nuclear power station between 1972 and 1996. The number of eggs deposited by each cohort at age 6 years was determined as follows (Aprahamian & Aprahamian, 2001):
where T = mean daily water temperature (° C) between June and August inclusive.
Three temperature scenarios were modelled: baseline, climate scenario 1 and climate scenario 2. The baseline scenario, reflective of present day conditions, incorporated mean ±s.d. June to August water temperature data of 17·8 ± 1·0° C between 1972 and 1996. Climate scenarios were based on UKCP09 medium emission predictions (Defra, 2010) and adopted a 1 and 2° C average temperature increase for scenarios 1 (18·8 ± 1·0° C) and 2 (19·8 ± 1·0° C), respectively. As the variation in mean temperature cannot be predicted in climate scenarios, the variation surrounding mean baseline data was applied. For the purposes of the model, temperatures were independently modelled for time step (i.e. t = 1 year). As follows: , where Tt = mean daily water June to August temperature in year t.
The total number of eggs deposited over the lifetime of a particular cohort (Y ) in year t was estimated as follows:
Baseline and climate scenario models were run in conjunction with Markov Chain Monte-Carlo (MCMC) simulations (PopTools; www.cse.csiro.au/poptools). Mean, variance and 95% c.i. were derived for each scenario and turbine mortality combination using 1000 iterations.
Water temperature data were obtained from a variety of Environment Agency (www.environment-agency.gov.uk) sources and the mean June to August temperature was calculated for the most downstream site in a river system. Time periods differed among rivers, as such direct comparison between systems was not possible. The data were used to provide an approximation as to the magnitude of the shift in latitude of the critical temperature needed for the persistence of A. fallax as a result of a 1 and 2° C warming scenario.
The relationship between stock, measured as the number of eggs deposited in year n and the number of recruits measured as the number of eggs produced by females age 6 years in year n + 6 [Fig. 2(a)] and when the estimate of recruits has been standardized using temperature as an explanatory variable [Fig. 2(b)], indicates that there is a weak density-dependent relationship and that stock explains a small proportion of the variability in recruitment measured 6 years later (r2 = 0·132, P > 0·05; with lower and upper 95% c.i. for a of 0·146 to 0·610 and for b of −4·181 × 10−7 to −9·977 × 10−8, respectively).
It is evident from Fig. 2 that the stock does not produce enough recruits from a single age-class spawning once to enable the population to persist.
Population persistence in relation to temperature
Alosa fallax in the Severn Estuary are multiple spawners and summation of the total number of eggs produced by a particular cohort over its lifetime indicates that in years when temperatures are elevated, a greater number of eggs were produced (Ey) than were deposited (Et) (Fig. 3) (r2 = 0·674, P < 0·01).
The relationship indicated that the equilibrium temperature is in the region of 17·4° C.
Effect of temperature on lifetime fecundity
A temperature component was fitted to the stock–recruitment relationship (equation 3) and the effect on lifetime fecundity computed for baseline and UKCP09 climate scenarios (Fig. 4). The model indicates that at a mean baseline temperature of 17·8° C, the population is able to persist whereby replacement is achieved (Fig. 4). Annual recruitment success, however, is highly variable among years, with population persistence not possible in colder years. Under climate scenario 1, the model suggests that an increase in water temperature of 1° C would be sufficient to support population persistence in all but the coldest years and result in on average, a 3·6 fold increase in the population (Fig. 4); a predicted increase in line with observed data (Fig. 3).
Under climate scenario 2, the model indicates that a 2° C increase in average water temperature to 19·8° C would see the population increase by on average, a factor of 6·3 and recruitment is predicted to be successful in all years (Fig. 4).
Effect of temperature on timing of freshwater phase of spawning migration
Aprahamian (1988) reported that peak immigration into the estuary occurred at temperatures ranging between 10·6 and 12·3° C, at between 119 and 134 days after 1 January (Fig. 5). An increase in water temperature of 1 and 2° C would result in the peak of the run occurring 6–10 days and between 16 and 17 days earlier, respectively (Fig. 5).
Effect of temperature on distribution
A long-term average mean June to August daily water temperature in excess of 17·5° C is suggested for a population of A. fallax to persist (Figs 3 and 4). Data on river temperature from the lower reaches were limited with little consistency in the time period available among river systems. Nevertheless, the data suggest that an increase in average water summer temperature of 1° C would result in habitat becoming suitable in terms of the temperature regime in large U.K. west coast rivers such as the River Dee (53·189° N) (15·9° C, 1979–1999) and the River Eden (54·972° N) (16·5° C, 1997–2007). For some of the small flashy (a river that responds quickly to an increase in rainfall) rivers intermediate in latitude between the Eden and the Dee, e.g. the River Kent (54·242° N; 15·9° C, 2005–2007), a 2° C rise in water temperature would be required to ensure a suitable thermal regime.
On the U.K. east coast, the Rivers Thames (latitude 51·477° N) and Ouse (53·697° N) (18·3° C, 1979–1999; 17·8° C, 1999–2009, respectively) already have a thermal regime similar to that of the River Severn. Even a 2° C rise would not extend suitable habitat as far north as the River Tyne (54·989° N) (15·0° C, 1995–2009).
Effect of additional mortality: persistence v. extinction
A tidal power barrage has the potential to affect (1) fish spawning for the first time [includes juveniles (age 0+ year) on their out migration in their first autumn and adults on their return to the river to spawn for the first time at between 2 and 6 years old (Aprahamian & Aprahamian, 2001)] and (2) repeat spawners, currently c. 50% of the spawning population, are repeat spawners with fish making up to seven spawnings (Aprahamian et al., 2003).
The effect of additional mortality prior to spawning for the first time suggests that an average increase of 50% (juvenile) and 90% (adults) on current mortality rates would result in the population becoming extinct under the current temperature regime (Fig. 6). The natural fluctuations in climatic conditions in any given year, however, result in population extinction occurring when turbine mortality rates are between zero and 60% and, zero and >90% for juvenile and adult A. fallax (based on 95% c.i.), respectively (Fig. 6). If average summer water temperatures were to increase by 1° C (18·8° C) or 2° C (19·8° C), then the estimated average mortality rates required for population extinction are c. 60 and 70% (juvenile) and c. 95% (adults) for each climate scenario, respectively (Fig. 6). Using a precautionary approach and adopting the worst case for both climate scenarios (lower 95% c.i.), the model indicates that maximum additional turbine mortality should not exceed 10 and 50% (juvenile) and, 10 and 90% (adults) for 18·8 and 19·8° C climate scenarios (Fig. 6).
In order for A. fallax to persist, a long-term summer average water temperature >17·5° C is needed. In regions northward of the Severn Estuary, river temperatures seldom exceed this threshold delineating the northern distribution of A. fallax. Under the predicted climate scenarios, the environment is predicted to warm sufficiently to increase the northern distribution range of A. fallax in the U.K. A 1° C rise in temperature is predicted to increase the northern range of A. fallax by 150 km beyond its existing northerly limit and possibly up to 350 km could be achieved. Such an increase is in line with predictions for other species (Hughes, 2000). The extent of any range increase, however, is likely to be constrained by local factors such as river size (volume) and available habitat.
Alosa fallax is regarded as a Lusitanian (warm temperate) species, and abundance is predicted to increase with a 1–2° C rise in temperature by a factor of 3·6 and 6·3, respectively. The mechanism by which temperature may be acting to control the population is likely to be through its effect on hatching success and growth rate. For Alosa alosa (L.), temperatures <16° C result in larval mortality at the egg stage (Hoestlandt, 1958). At temperatures between 16 and 18° C, larval condition is poor resulting in difficulty emerging from the egg (Cassou-Leins & Cassou-Leins, 1981), optimum temperature for the survival of eggs and larvae being in the region of 20° C (K. Charles & P. Jatteau, unpubl. data). When temperatures increase, growth may be faster, such that 0+ year fish are vulnerable to predation from aquatic invertebrates for a shorter period of time (Mann, 1991) and thus, enhancing population persistence.
Temperature may also affect food production, with primary production greater in warmer years. Year-class strength in A. sapidissima has been shown to be positively correlated with zooplankton density, as shown by an increase in the percentage of larval fish (<13 mm) with food in their guts (Crecco & Savoy, 1987). The similarity in the coefficient of variation in year-class strength between the study of Aprahamian & Aprahamian (2001) and that of Henderson & Seaby (1999) supports the conclusion that the mechanism operates during the first 4 months before the fish migrate seaward in the autumn (Claridge & Gardner, 1978; Aprahamian, 1988).
The weak relationship between stock size and recruitment suggests that the population is regulated mainly through abiotic processes, specifically temperature. Similar reports have been made for A. sapidissima on the Connecticut River, U.S.A. (Crecco et al., 1986) and for the landlocked population of A. pseudoharengus in Lake Huron, North America (Henderson & Brown, 1985). In these studies, density-dependent factors of recruitment without the influence of climatic factors removed, accounted for between 5 and 7% of the variation in recruitment in Lake Huron (Henderson & Brown, 1985) and c. 2% on the Connecticut River, U.S.A. (Crecco et al., 1986). In contrast, strong density-dependent regulation is evident in semelparous populations of A. alosa in the River Loire, France (C. Mennesson-Boisneau, unpubl. data) and in the Gironde, France (Martin-Vandembulcke, 1999).
There is at present little information on the levels of turbine-induced mortality that might be experienced by A. fallax in a tidal environment. Several factors arising from passage through a barrage have been attributed to juvenile and adult alosine mortality and include pressure effects, blade strike and shear effects (Dadswell & Rulifson, 1994). Dadswell & Rulifson (1994) citing the unpublished studies of W. E. Hogans & G. D. Melvin reported mean ± 95% c.i. mortality levels of 46·3 ± 34·7% and 21·3 ± 15·2%, at a tidal turbine in the Annapolis Estuary, Canada, during 1985 and 1986, for adult A. sapidissima. The difference in mortality rate between the two studies was attributed to an increase in turbine efficiency and reduced handling stress (Dadswell & Rulifson, 1994).
Subsequently, Dadswell & Rulifson (1994) and Gibson & Myers (2002) used net capture rates to investigate the juvenile mortality (0+ year juveniles). Dadswell & Rulifson (1994) reported an overall mortality rate of 54·4% for A. sapidissima. Trials in 1999 assessing turbine mortality for age 0+ year A. sapidissima, Alosa aestivalis (Mitchill) and A. pseudoharengus, however, indicated overall mean mortality rates (±95% c.i.) of 23·4% (6·1–58·8), 8·1% (3·5–17·2) and 7·7% (1·5–31·4), respectively, for a single passage (Gibson & Myers, 2002). Differences in turbine mortality estimates between studies were attributed to the duration the nets were deployed (Gibson & Myers, 2002).
Unpublished experimental studies have been used to predict injury rates for juvenile (70 mm) alosines during a Severn tidal power generating cycle. Overall mortality was estimated at 53%; the majority of which was attributed to shear effects (48%) and the remaining mortality to blade strike (4·9%) (A. Turnpenny, unpubl. data). In contrast, a study at Annapolis Royal tidal power project (Nova Scotia, Canada) suggested that the main injuries to juvenile alosines were due to pressure effects (54·4%), with shear (1·2%) and mechanical effects (3·4%) secondary; 41% had no signs of damage (Stokesbury & Dadswell, 1991). It is unclear as to the exact mechanism that causes juvenile mortality but, irrespective of the mechanism, it is clear that turbine mortality has the potential to greatly affect fish survival during passage through a tidal barrage turbine.
In estuaries, fishes move in and out with the tide, so there is the possibility that they may make repeated passages through the turbines, which is in contrast with in-river structures where a single passage is assumed. Nothing is known about the routes taken by the adult and juvenile A. fallax in the outer Severn Estuary. Variation in catch between flood and ebb tide suggests that prespawning adults tend to use the flood tide in the main channel to facilitate upstream progress, and the ebb tide to reduce displacement from the stem. A movement inshore would reduce the risk of passing through generating turbines, assuming that fish pass facilities were placed close to the shore. Postspawning fish are not, however, caught in any number in the fishery which operates close to the bank, indicating that seaward movement may be facilitated by swimming in the main channel on the ebb tide.
Aprahamian (1988) showed the consequence of additional mortality on the mature stock of the Severn A. fallax population to be a decline in mean mass and in the number of spawnings. The indications from this study are that the populations in the River Severn and presumably also in the Rivers Wye, Usk and Tywi are barely able to sustain themselves under current climate conditions and an iteroparous life history is essential for A. fallax to maintain a self-supporting population. Leggett (1976, 1977) and Gibson & Myers (2003) concluded a similar requirement for A. sapidissima and A. pseudoharengus populations close to their northern limits in North America. If 0+ year A. fallax are as vulnerable to turbine mortality as A. sapidissima with a mortality rate of between 23·4 and 46·3% for a single passage, then an increase in mortality of c. 30% is likely. At this level of mortality, it is predicted that the population would no longer be able to persist, even if climate temperatures were to increase by 1° C. A warming of 2° C would enable A. fallax to sustain the additional mortality from a single passage. Although the population would be larger and more resilient due to greater annual recruitment success, it would effectively only take two passages through the turbine before the population was no longer self-sustaining. If A. fallax responds in a similar way to A. aestivalis and A. pseudoharengus where turbine mortality rates are considerably lower, or if turbine mortality can be reduced through deployment of a more ‘fish-friendly’ turbine (Hecker & Cook, 2005), then it is possible that the greater number of A. fallax will successfully pass through the turbines to support population persistence.
The increase in temperature that is forecasted to occur as a result of global warming is predicted to benefit the populations of A. fallax in the U.K. The consequence of a tidal barrage across the estuary of the River Severn, to reduce the production of greenhouse gases will, however, most likely eliminate a self-sustaining population in this river. The effects on the populations in the Rivers Wye and Usk would be dependent on the specific scheme chosen. The increase in temperature will mean that rivers further north may have an acceptable thermal regime which would enable a population of A. fallax to persist, should a sufficient number of A. fallax disperse to those areas. To offset the possible effects of a tidal barrage, the rate of warming would ideally need to proceed with the development of a barrage in order for populations to develop, although founder fish originating from French and Irish populations may establish new populations. Certainly, A. fallax has been reported as far north as the Solway Firth (54·888° N), but as yet there is no evidence of a spawning population in its rivers (Aprahamian & Aprahamian, 1990). Thus, a paradox exists between trying to reduce the effects of greenhouse gas emissions and the possible local extinction of an internationally considered rare and endangered species.
We thank S. M. Lester and N. Mott for their very valuable assistance and gratefully appreciate the extremely constructive reviews of two anonymous referees and I. J. Winfield. The views expressed in this paper are those of the authors and not necessarily those of the Environment Agency or APEM Ltd.