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Stable abundance, but changing size structure in grenadier fishes (Macrouridae) over a decade (1998–2008) in which deepwater fisheries became regulated
ARTICLE IN PRESS Deep-Sea Research I 57 (2010) 434–440
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Stable abundance, but changing size structure in grenadier fishes (Macrouridae) over a decade (1998–2008) in which deepwater fisheries became regulated Francis Neat , Finlay Burns Marine Scotland-Science, Marine Laboratory, P.O. Box 101, 375 Victoria Road, Aberdeen AB11 9DB, UK
a r t i c l e in f o
a b s t r a c t
Article history: Received 7 October 2009 Received in revised form 17 December 2009 Accepted 27 December 2009 Available online 7 January 2010
A ten-year time series (1998–2008) from a trawl survey of the continental slope of the NE Atlantic was analyzed to assess temporal variation in the abundance and length frequency of seven species of deepwater grenadier fish. This period coincided (in 2003) with the regulation of deepwater fisheries in this area. None of the species declined in numbers or biomass over the period, and 2 species significantly increased. This suggests that the declines in abundance of these deepwater species following the onset of fishing in the 1970s may now have stabilized, albeit at much lower levels than the virgin biomass. Although two metrics of body size (mean length and maximum length) did not show any evidence for consistent decrease over time, there were significant changes in the overall length–frequency distributions. The species found in shallower depths (500 m) had a greater number of larger individuals in 2008 whereas those found deeper (1500 m) tended to have a greater number of smaller individuals. This suggests the presence of a lagged indirect effect of fishing on species that live beyond the actual depths that fishing takes place. & 2010 Elsevier Ltd. All rights reserved.
Keywords: Deepwater fisheries Coryphaenoides CPUE Length frequency Time series
1. Introduction The continental slopes of the world’s oceans separate the shelf seas from the abyssal plains and are characterized by steep gradients in seabed conditions and enhanced productivity (Gage and Tyler, 1993). Although slope environments represent only a small fraction of the area of the oceans, they support large and diverse fish assemblages. Consequently they are the focus of the majority of deepwater trawl fisheries. Trawlers began fishing the slopes of the NE Atlantic in the1970s where they operated as deep as 1500 m, although most effort was at depths around 1000 m where commercially valuable species such as roundnose grenadier (Coryphaenoides rupestris), orange roughy (Hoplostethus atlanticus), blue ling (Molva dipterygia), black scabbard (Aphanopus carbo) and deepwater sharks (e.g. Centrophorus squamosus) were most abundant (Gordon, 2001; Gordon et al., 2003). The fish assemblage in this area was also well studied in the 1970s just prior to the establishment of the major commercial deepwater trawl fisheries (Gordon, 1986). As research progressed it became apparent that many deepwater fish attained very old ages, had slow growth rates and were very late to mature, the combination of which made them unproductive and highly vulnerable to
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over-exploitation (Bergstad, 1995; Koslow et al., 2000; Morato et al., 2006). The example often cited is that of orange roughy (Clark, 2001) which is among the most vulnerable of all deepwater fishes because of its slow growth, late age at maturity (Minto and Nolan, 2006), extreme longevity (Andrews et al., 2009) and aggregating behavior. Furthermore, since the by-catch taken by deepwater fisheries is often large (Allain et al., 2003) and suffers total mortality due to the barotrauma of being brought from great depths, it became apparent that the entire fish assemblage of the slope could be highly vulnerable. Such concerns seem to have been borne out, not only by the decline in catches (Lorance and Dupouy, 2001), but also by two recent studies based on fisheries-independent data. The first, an analysis of data from the western North Atlantic suggested dramatic declines (up to 90%) in abundance of several deepwater slope species since the onset of deepwater fishing (Devine et al., 2006). The second, an analysis of the numerical abundance of the deepwater fish from the slope of the Porcupine bank area in the NE Atlantic suggested that by the 1990s numbers had declined to approximately half of what was estimated prior to exploitation (Bailey et al., 2009). Declines in abundance are, however, inevitable population consequences of exploitation (Haedrich and Barnes, 1997; Pauly et al., 1998; Hilborn et al., 2003), and theoretically a biomass reduction of half may be required to attain maximum productivity (Schnute and Richards, 2002). The declines in deepwater species were
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certainly dramatic; however, given that the absolute biomass and population dynamics of the fishery were largely unknown, it is difficult to determine whether this is indicative of terminal depletion or just the expected trajectory of a fishery on its way to reach maximum productivity. If the former, it would be expected that the stocks would continue to decline to commercial extinction. On the other hand, it could be argued the perception of the severity of the decline is due to the fact that unlike most shallow-water fisheries that have faced centuries of exploitation, the declines in deepwater stocks can be referenced to data that were collected when the stock was in a recently unexploited, ‘virgin’ state. If so, and the decline reflects a particularly fast case of ‘the cropping of the stock’, then it would be predicted that the stocks eventually stabilize at a lower absolute biomass. These two alternative interpretations need not be mutually exclusive and may both underlie observed patterns in deepwater fisheries. The problem is teasing the two apart and this has been hampered until now by a lack of fisheries-independent information from the years that followed the initial intensive exploitation. As the deepwater fisheries grew unregulated, so did concerns over their sustainability. In 2003, a total allowable catch (TAC) system was implemented for managing the ten major deepwater stocks in the NE Atlantic. This effectively capped any further expansion of deepwater fisheries until such time that there was good evidence that any increase could be sustainable. The International Council for the Exploration of Sea (ICES) held the view that most stocks were ‘probably outside safe biological limits’ and advised on precautionary fishing effort reductions. Since 2005 TACs and subsequently fishing effort and landings of most deepwater species have declined (ICES, 2008). For example, the TAC for roundnose grenadier decreased from over 5000 tonnes to under 4000 tonnes with landings declining to under 3000 tonnes in 2008 (Lorance et al., 2008; ICES, 2008). The consequences of such regulations have had for the populations of slope-dwelling fish remained largely unknown. There is therefore a pressing need for information from the period following the initial exploitation and over the period during which the fisheries were regulated, i.e. the past decade. Since 1998 Marine Scotland (formerly Fisheries Research Services) has undertaken a deepwater trawl survey of the continental slope to the west of Scotland, the data of which can be used to address these issues. Here we estimate recent trends in the abundance and size structure of one family of slope-dwelling fishes, the grenadiers (family: Macrouridae). Only the roundnose grenadier is commercially targeted as part of a mixed trawl fishery, but the six other species may be taken as by-catch. Grenadiers were chosen as the focus for several reasons: (1) they are numerous enough to obtain good estimates of abundance and length frequencies, (2) The seven species commonly found span the full extent of the depth gradient of the continental slope in this region (Lorance et al., 2008). Therefore some species are within depth strata directly exposed to fishing while others are beyond the depth of current fishing limits ( 1200 m), (3) they are one of the better studied families of deep-sea fishes (Moranta et al., 2007; Orlov and Iwamoto, 2009) and so information on life-histories is available to aid interpretation of results. Overall therefore, this family of fishes is a good model group for investigating fishing impacts on the continental slope. We first address the question of whether there is any evidence of a change over the past decade in the abundance of the grenadiers in terms of numbers and biomass. Secondly because exploitation often has consequences for the size structure of populations, we ask if there is any evidence of a change over the past decade in length– frequency distribution and body size.
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2. Materials and methods 2.1. The Marine Scotland deepwater survey Marine Scotland used FRV Scotia to undertake a deepwater bottom trawl survey of the continental slope in the years 1998, 2000, 2002, 2004, 2005, 2006, 2007, 2008. The survey took place in September of each year and trawling was carried out mainly during daylight hours with a typical duration of 2 h bottom time. The full survey area (located within ICES area VIa) stretched along and down the slope from 551 to 591N at approximately 91W and covered depths from 300 to 1900 m. This full dataset was used to investigate the depth distribution patterns of the different Macrourid species. However, not all of this area was consistently surveyed at all depth strata and all years. Therefore for the purposes of modeling abundance and length frequency, data were derived from hauls taken from an area that spanned 56–58.51N at approximately 91W and at depths of between 400 and 1600 m (Fig. 1). This area covers five ICES statistical rectangles (41EO, 42EO, 43EO, 44EO, 45EO). A total of 110 hauls were available for analysis which gave near complete coverage in terms of ICES squares, depth strata (500, 1000, 1500 m) and years, the only exception being 1500 m in the year 1998 (Table 1). Trawling depth was kept to 7100 m of each depth stratum, but typically this did not vary more than 50 m within any particular haul.
2.2. Trawl gear specifications A bottom trawl (BT184) was rigged with 2100 rock-hopper ground gear (Jackson Trawls Ltd, Peterhead, UK), 1700 kg doors of area 5.82 m2 (Morgere, St.Malo, France), 100 m sweeps and floats rated to 2500 m. Warp to bottom-depth ratio ranged between
Fig. 1. A map of the survey area showing depth shading with 500, 1000 and 1500 m depth contours (white), the ICES statistical rectangle numbers and trawling positions (black lines). The island of St. Kilda (UK) is shown for reference.
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3. Results
Table 1 Numbers of hauls available per depth category per year for modeling. Year
500 m
1000 m
1500 m
Total
1998 2000 2002 2004 2005 2006 2007 2008 Total
5 6 4 6 4 5 5 5 40
5 5 4 4 4 5 5 5 38
NA 5 5 4 4 5 5 5 33
10 16 13 14 12 15 15 15 110
approximately 2.5:1 and 2:1 decreasing gradually with depth. The cod-end was fitted with an internal liner with 20 mm mesh size which ensured retention of fishes as small as a few cm in length. The entire catch was sorted and identified to species level. For each haul, the total weight for each species was recorded. Due to the tendency for the fragile tails of Macrourid species to break off in the process of being trawled, total length was not a suitable measure; instead length was measured as the distance (cm) from the snout to the first ray of the anal fin (the pre-anal fin length—PAFL). The total number of individuals for each species was determined by measuring the PAFL for each individual fish and creating a length–frequency distribution. There were instances where a species was too numerous for all fish to be measured and a subsample by weight was taken with the final numbers per cm size class raised accordingly. Efforts were made to ensure this subsample was representative of the entire catch, i.e. baskets were sampled randomly with respect to the sequence by which they were sorted and weighed. 2.3. Data and statistical analysis Catch per unit effort (CPUE) data were analyzed using generalized linear models (GLM). Two measures of CPUE were considered: numbers per hour (N h 1) of trawling and kilograms per hour (kg h 1) of trawling. To account for the over-dispersed nature of the N h 1 data the GLM was fitted using a negative binomial distribution. For the kg h 1 data the GLM was fitted with a gamma distribution. Although the time series spanned 10 years, this was of insufficient length to formally test for autocorrelation within the data. Instead the residuals of the models were inspected with respect to year to see if there were any indications of cyclic temporal trends. Similarly, to assess the degree of spatial autocorrelation the model residuals were examined with respect to sequential ICES statistical rectangles. Year, depth category and ICES squares (effectively latitude) were treated as factors in the models, and if interactions between factors were not significant the interaction terms were removed. In most cases there was one haul for each year, depth and ICES square, although some missing values were inevitable due to invalid hauls (e.g. net getting stuck and/or being damaged on the seabed) or constraints (e.g. weather) on completing all hauls in some years. For each species only hauls from their core depth distribution (see Section 3.1) were used. For analysis of mean and maximum length (per haul) a linear regression was used to model mean length or maximum length from each haul. The Kolmogorov–Smirnoff (K–S) test was used to determine if the length–frequency distribution of the first year of the time series was significantly different from the last. Any consistent trends in the direction of distribution shift over years were taken to be indicative of a temporal pattern (a shift to left indicates a greater proportion of smaller fish and vice versa) whereas a lack of consistency over years was taken to represent stochastic variability.
3.1. Species depth distributions and assignment of core depth strata Although 10 species of grenadier have been recorded on the survey, only 7 species are encountered in sufficient frequency to model temporal and spatial patterns effectively. These were Caelorhyncus caelorhyncus, Nezumia aequalis, C. rupestris, Coryphaenoides guentheri, Coryphaenoides mediterraneus, Caelorhyncus labiatus and Trachyrinchus murrayi. There were clear patterns of abundance (square root N h 1) with respect to depth (Fig. 2). The distribution of C. caelorhyncus ranged from o300–1000 m, with 82% of individuals caught between 400 and 600 m. Its core depth stratum was taken to be 500 m. N. aequalis was found deeper between 400 and 1300 m, with 95% of individuals caught between 700 and 1100 m. Its core depth stratum was taken to be 1000 m. C. rupestris had the widest depth distribution ranging from 500 to 41800 m with 91% of individuals caught between 900 and 1600 m. To account for this both 1000 and 1500 m were taken to reflect its core distribution. C. guentheri was found between 800 and 1800 m with 98% of individuals caught between 1300 and 1800 m. Its core depth stratum was taken to be 1500 m. C. mediterraneus was found between 1200 and 1800 m with peak numbers occurring between 1300 and 1800 m. Its core depth stratum was taken to be 1500 m. C. labitaus was found between 800 and 1800 m with 99% of individuals caught between 1300 and 1800 m. Its core depth stratum was taken to be 1500 m. T. murrayi was found between 800 and 1600 m with 91% of individuals caught between 1200 and 1600 m. Its core depth stratum was taken to be 1500 m. These observations are highly comparable to data presented by Lorance et al. (2008; Table 1) for data collected from the same area in the 1970s and 1980s suggesting the bathymetric distributions of the species are stable over decadal time spans. 3.2. Variation in relative abundance, mean size and length frequency For C. caelorynchus and N. aequalis data was available for all years in the time series, but the other species had core depth ranges that extend deeper than 1000 m, data was available for all years except 1998. The statistical results of the GLM and linear model analyses are presented in Table 2. 3.2.1. C. caelorynchus (500 m) Both numbers per hour (N h 1) and kilograms per hour (kg h 1) increased significantly between 1998 and 2008 (Fig. 3a). There was no effect of ICES square and no significant interaction. There was no trend over time with respect to mean length, although there was significant variability between ICES squares. Maximum length however did show a significant increase over years (Fig. 3b). The length–frequency distribution in 1998 was significantly different from that in 2008 (K–S test, D=0.64, P o0.001) with a higher frequency of larger individuals in 2008. This pattern also held for 2000 and 2002 suggesting a directional trend from the early to later years (Fig. 3c). 3.2.2. N. aequalis (1000 m) Both N h 1 and kg h 1 showed a trend for increasing numbers in later years, although this was only significant for kg h 1 reflecting a sudden increase in 2004 that remained high thereafter (Fig. 4a). There was also a significant effect of ICES square with greater numbers and biomass in more southerly squares. There was no significant trend over time with respect to mean or maximum length (Fig. 4b) and no significant variability between ICES squares. The length–frequency distribution of this species in
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70
C. caelorinchus
Sqrt Abundance (N.h-1)
60
N. aequalis 50
C. rupestris 40
C. guentheri 30
C. mediterraneus 20
C. labiatus 10
T. murrayi 0 300 400 500 600 700 800 900 1000 1100 1300 1500 1700 1800 Depth (m) Fig. 2. Bathyal distribution of the 7 species of grenadier expressed as square root abundance and standardized to 1 h. of trawling. Note no hauls were taken at depths less than 300 m.
Table 2 Results of GLM analyses of CPUE and size metrics for each species. Species depth (m) number of hauls
Variable
Year and trend
ICES Square
Depth
C. caelorynchus 500 m n = 38
N h1 kg h 1 Mean length Max length
Po 0.001 (m) Po 0.001 (m) NS, P= 0.11 (–) Po 0.01 (m)
NS, P =0.11 NS, P =0.42 P o 0.01 P o0.01
N/A N/A N/A N/A
N. aequalis 1000 m n = 36
N h1 kg h 1 Mean length Max length
NS, P= 0.07 (m) P= 0.01 (m) NS, P= 0.47 (–) NS, P= 0.57 (–)
P o0.001 P =0.015 NS, P =0.43 NS, P =0.39
N/A N/A N/A N/A
C. rupestris 1000–1500 m n = 65
N h1 kg h 1 Mean length Max length
NS, P= 0.32 (–) NS, P= 0.13 (–) NS, P= 0.89 (–) Po 0.001 (m)
P o0.05 P =0.012 P =0.012 P o 0.01
P o0.001 P o0.001 P o0.001 P o0.05
C. guentheri 1500 m n = 33
N h1 kg h 1 Mean length Max length
NS, NS, NS, NS,
(–) (–) (–) (–)
P =0.001 NS, P =0.39 P =0.002 NS, P =0.09
N/A N/A N/A N/A
C. mediterraneus 1500 m n = 33
N h1 kg h 1 Mean length Max length
Po 0.001, (m) NS, P= 0.08 (m) NS, P= 0.12 NS, Po 0.17
P =0.001 P =0.015 NS, P =0.16 NS, P =0.49
N/A N/A N/A N/A
C. labiatus 1500 m n = 33
N h1 kg h 1 Mean length Max length
Po 0.001 (m) NS, P= 0.44 (–) Po 0.001 (k) P= 0.03 (-)
P o0.001 P =0.019 P =0.72 NS, P =0.78
N/A N/A N/A N/A
T. murrayi 1500 m n = 33
N h1 kg h 1 Mean length Max length
NS, NS, NS, NS,
P o0.001 P =0.019 NS, P =0.69 NS, P =0.69
N/A N/A N/A N/A
P= 0.56 P= 0.17 P= 0.21 P= 0.33
P= 0.28 (–) P= 0.41 (–) P= 0.31 (–) Po 0.39 (–)
N/A = not applicable, P =P-value, NS =non-significant, (m) = increasing trend, (–) = no trend, (k) = decreasing trend.
1998 was significantly different from that in 2008 (K–S test, D= 0.23, Po0.001; Fig. 4c) with larger individuals more frequent in 2008, however, this pattern was not consistent across other years and may reflect short-term stochastic variation as opposed to a longer term trend.
3.2.3. C. rupestris (1000–1500 m) For both N h 1 and kg h 1 (Fig. 5a) there was no effect of year. There was a significant effect of depth with greater numbers and biomass at 1500 m and there was significant variability between ICES square, although these differences were not consistently
related to latitude. There appears to have been no significant change in the mean length of this species over years. Mean and maximum length was significantly greater at 1500 m and varied significantly between ICES squares. The maximum length of this species showed a small but significant increase over the time period (Fig. 5b), although the presence of a significant interaction between depth and year suggested the increase was apparent only at 1500 m. The cumulative length–frequency distribution of this species in 2000 was significantly different from that in 2008 (K–S test, D= 0.42, Po0.001; Fig. 5c). This appears to be related mainly to a decrease in the number of small individuals ( o9 cm) in later years.
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Figs. 3–9. Indices of abundance, size metrics and length–frequency data for each species in each year: C. caelorhyncus (Fig. 3), N. aequalis (Fig. 4), C. rupestris (Fig. 5), C. guentheri (Fig. 6), C. mediterranaeus (Fig. 7), C. labiatus (Fig. 8) and T. murrayi (Fig. 9). For each species: (a) Mean + - 95% cls CPUE data (N h 1—filled circles, solid line and kg h 1open circles, dashed line); (b) mean PAFL (light gray bars, plus standard deviation) and average maximum PAFL (dark gray bars, plus upper range) and (c) cumulative percentage of individuals for each length class. Solid black curve = 1998 or 2000, dashed curve =2008, gray curves =intervening years.
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3.2.4. C. guentheri (1500 m) For neither N h 1 nor kg h 1 (Fig. 6a) was there any significant effect of year. There were however significant effects of ICES square with numbers steadily increasing from south to north. There was no significant trend over time with respect to mean or maximum length (Fig. 6b), although there was significant variability between ICES squares. The cumulative frequency distribution of this species in 2000 was significantly different from that in 2008 (K–S test, D= 0.08, Po0.05; Fig. 6c) with higher frequencies of smaller individuals in 2008. 3.2.5. C. mediterraneus (1500 m) For both N h 1 and kg h 1 (Fig. 7a) there was an increasing trend over time, although this was only significant for N h 1. There was a significant effect of ICES square (P= 0.001) with numbers and biomass steadily increasing from south to north. There was no significant trend over time with respect to mean or maximum length (Fig. 7b) and no significant variability between ICES squares. The cumulative frequency distribution of this species in 2000 was significantly different from that in 2008 (K– S test, D=0.27, Po0.001; Fig. 7c) with smaller individuals more frequent in 2008. This pattern was consistent in each subsequent year. 3.2.6. C. labiatus (1500 m) For N h 1 (Fig. 8a) there was a significant effect of year with numbers increasing until 2008, which showed a downturn. There also was a significant effect of ICES square with numbers steadily increasing from south to north. For kg h 1 (Fig. 8a) there was no significant effect of year, but there was a greater biomass in more northerly squares. There was a significant effect of year with respect to mean length (Fig. 8b) with a consistent decline in mean size since 2002 and no significant variability between ICES squares. There was also a significant effect of year with respect to maximum size, although no consistent trend over years was apparent. The cumulative frequency distribution of this species in 2000 was significantly different from that in 2008 (K–S test, D= 0.30, Po 0.001; Fig. 8c) with smaller individuals more frequent in 2008. The pattern was not however consistent from year to year. 3.2.7. T. murrayi (1500 m) For neither N h 1 nor kg h 1 (Fig. 9a) was there any effect of year. There was an effect of ICES square with numbers and biomass steadily increasing from north to south. There was no significant trend over time with respect to mean or maximum length (Fig. 9b) and no significant variability between ICES squares. The cumulative frequency distribution of this species in 2000 was not significantly different from that in 2008 (K–S test, D= 0.25, NS; Fig. 9c) and showed little difference across all years.
4. Discussion Most fish stocks will show an initial downturn in abundance following exploitation (Schnute and Richards, 2002); this is not unique to deepwater fisheries. What is unique to deepwater fisheries is that only 40 years ago they were in a virgin biological state, and that they underwent a particularly fast decline immediately following the onset of fishing (Devine et al., 2006; Bailey et al., 2009). Deepwater fish differ from most shallowwater fish in being very long-lived, slow growing and late maturing (Koslow et al., 2000). Consequently they are less productive and likely to be highly vulnerable to over-exploitation (Morato et al., 2006). Much concern and debate has been raised
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over stock declines and whether they can be sustainably exploited (Roberts, 2002). Some deepwater species such as orange roughy (Clark, 2001) almost certainly cannot be fished sustainably because catches would have to be so low as to be unprofitable. For other deepwater species, however, biological and economic factors may be rather different. Provided fisheries for such species are carefully regulated on the basis of good data there may be sustainable levels of profitable exploitation. The results from this study demonstrate that the grenadiers from the NE Atlantic have not declined in the past ten years. Although the current area studied is some 200 miles north of the study area of Bailey et al. (2009), the slope-dwelling fishes would most likely have undergone a similar decline because the main commercial trawl fisheries extended along most of the slope (Gordon et al., 2003). Assuming this to be the case, it would appear that the decline may have now stabilized. The stocks, however, show little sign of recovering, and in absolute terms they are probably still a fraction of what they were at of preexploitation virgin biomass levels. This is perhaps not surprising given that estimates of stock recovery time for C. rupestris range from 14 to 80 years (Baker et al., 2009). Thus while the rate of population decline caused by fishing appears to have been particularly fast in these deepwater species the fact that they do appear to be stabilizing suggests they are not in terminal decline. Unfortunately it is difficult to compare directly our abundance estimates with those collected in the 1970s because of the differences between survey gear, research vessels and area. Nevertheless, as indices of relative abundance, the results are clear-cut and because the survey design and analyses controlled for or accounted for the effects of season, depth and sampling area, we are confident they reflect genuine temporal patterns. Furthermore it is unlikely that the present results have arisen from a shift in core depth distribution of each species over the time period in part because the depth distributions for these species reported here are highly comparable to these reported in the same area in the 1970s (Lorance et al., 2008). The relative stability since 1998 may reflect the introduction (in 2003) of a management regime including TACs for commercial deepwater species which included roundnose grenadier. Since 2005 TACs have generally declined and some directed fisheries, such as for orange roughy and deepwater sharks are being phased out altogether. Thus fishing pressure has been alleviated and these data cautiously suggest that it may have been sufficient to prevent the further decline of the grenadier fishes. It is interesting that those species that showed an increase in abundance were the shallower, smaller species such as C. caelorynchus and N. aequalis. This may be because of their lower age and size at maturity (Coggan et al., 2000) compared to species such as C. rupestris (Bergstad, 1990). Alternatively because these smaller species occur in the depth strata at which the most intensive trawling activity took place, they may have experienced a greater relative reduction in fishing pressure and the trend is thus more pronounced. Although the grenadiers are a very diverse family (Orlov and Iwamoto, 2009) with a broad range in body sizes, growth rates and longevity, their energetics and life-history are fundamentally different from the most vulnerable deepwater species that include seamount-associated species, such as orange roughy (Koslow, 1996) and deepwater sharks which have very ˜ o´n et al., 2006). Consequently the low reproductive rates (Ban present findings should not be taken as indicative that all deepwater species may be expected to show signs of abundance stabilization following the regulation of fisheries. With respect to size-based metrics there were no signs that the mean or maximum size of individuals had decreased in the past decade. These two metrics are often sensitive indicators of fish populations in process of being over-exploited (Haedrich and
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Barnes, 1997; Bianchi et al., 2000). Ten years may be insufficient time to observe such changes in long-lived species and it is perhaps more informative to look at the overall size structure of the populations. Here there were some interesting patterns over time particularly when all the species are compared across the full depth range of the slope. The one species (C. caelorynchus) with a core depth distribution on the upper slope at 500 m had a greater number of larger individuals in the population in 2008 than in 1998, 2000 and 2002, suggesting a directional trend. This trend was not evident in those species found at 1000 m (N. aequalis and C. rupestris) and was in the opposite direction in 3 out the 4 species found at 1500 m (C. guentheri, C. mediterraneus and C. labiatus). The increase in the number of larger individuals in the upper slope species may reflect the recent reduction in fishing mortality with individuals living longer and growing larger. On the lower slope, the increase in the number of smaller individuals into the populations may be related to what Bailey et al. (2009) describe as the impact of fishing extending well beyond the actual depth of fishing. Although Bailey et al. (2009) noticed such effects with respect to abundance, it is likely the size structure of the populations will also be affected. The mechanisms underlying such changes are not clear, but one possibility is an indirect effect that arises from the reduction of apex predators such as deepwater sharks (Jones et al., 2005) and blue ling (Lorance and Dupouy, 2001), which in turn has lead to an increase in the survival probability of small individuals.
5. Conclusion The data presented here suggest that the alarming declines in deepwater species reported after the fisheries began in the Rockall Trough may have stabilized in at least one major group of slope-dwelling fishes. It is cause for cautious optimism that the recent reduction and regulation of deepwater fishing in this area may have been sufficient to prevent further decline. There is, however, little sign of recovery to former levels and without more information on stock productivity (such as growth rates) it would be premature to comment on whether fishing for species such as roundnose grenadier could be considered sustainable.
Acknowledgments The research was supported by the Scottish Government (MF0763). We are especially grateful to FRV Scotia Fishing Masters Neville Ball and Peter Carmichael, their crew and the many Marine Laboratory staff that helped collect these data, notably Kevin Peach and Emma Jones. We further thank Peter Wright, Neil Campbell and Bill Turrell for comments on the paper, and M. Bacon, I. Priede and three anonymous referees for their critical and constructive review of the paper. References Allain, V., Biseau, A., Kergoat, B., 2003. Preliminary estimates of French deepwater fishery discards in the Northeast Atlantic Ocean. Fish. Res. 60, 185–192, doi:10.1016/S0165-7836(02)00031-0. Andrews, A., Tracey, D.M., Dunn, M.R., 2009. Lead-radium dating of orange roughy (Hoplostethus atlanticus): validation of a centenarian life span. Can. J. Fish Aquat. Sci. 66, 1130–1140, doi:10.1139/F09-059. Bailey, D.M., Collins, M.A., Gordon, J.D.M., Zuur, A.F., Priede, I.G., 2009. Long-term changes in deep-water fish populations in the northeast Atlantic: a deeper
reaching effect of fisheries? Proc. R. Soc. Lond. B 275, 1965–1969, doi:10.1098/ rspb.2009.0098. Baker, K.D., Devine, J.A., Haedrich, R.L., 2009. Deep-sea fishes in Canada’s Atlantic: population declines and predicted recovery times. Environ. Biol. Fishes 85, 79–88, doi:10.1007/s10641-009-9465-8. ˜ o´n, R., Pin ˜ eiro, C., Casa, M., 2006. Biological aspects of deep-water sharks Ban Centroscymnus coelolepis and Centrophorus squamosus in Galician waters (north-western Spain). J. Mar. Biol. Assoc. UK 86, 843–846, doi:10.1017/ S0025315406013774. Bergstad, O.A., 1990. Distribution, population structure, growth and reproduction of the roundnose grenadier Coryphaenoides rupestris (Pisces: Macrouridae) in the deep waters of the Skagerrak. Mar. Biol. 107, 25–39. Bergstad, O.A., 1995. Age determination of deep-water fishes: experiences, status and challenges for the future. In: Hopper, A.G. (Ed.), Deep-water Fisheries of the North Atlantic Oceanic Slope. Kluwer Academic Publishers, Amsterdam, pp. 267–283. Bianchi, G., Gislason, H., Graham, K., Hill, L., Jin, X., Koranteng, K., ManickchandHeileman, S., Paya, I., Sainsbury, K., Sanchez, F., Zwanenburg, K., 2000. Impact of fishing on size composition and diversity of demersal fish communities. ICES J. Mar. Sci. 57, 558–571, doi:10.1006/jmsc.2000.0727. Clark, M., 2001. Are deepwater fisheries sustainable? — the example of orange roughy (Hoplostethus atlanticus) in New Zealand. Fish. Res. 51, 123–135, doi:10.1016/S0165-7836(01)00240-5. Coggan, R.A., Gordon, J.D.M.., Merrett, N.R., 2000. Aspects of the biology of Nezumia aequalis from the continental slope west of the British Isles. J. Fish Biol. 54, 152–170, doi:10.1111/j.1095-8649.1999.tb00619.x. Devine, J.A., Baker, K.D., Haedrich, R.L., 2006. Fisheries: deep-sea fishes qualify as endangered. Nature 439, 29, doi:10.1038/439029a. Gage, J.D., Tyler, P.A., 1993. Deep-sea Biology: a Natural History of Organisms at the Deep-Sea Floor. CUP, Cambridge 504pp. Gordon, J.D.M., 1986. The fish populations of the Rockall Trough. Proc. Roy. Soc. Edin. 88B, 191–204. Gordon, J.D.M., 2001. Deep-water fisheries at the Atlantic Frontier. Cont. Shelf Res. 21, 987–1003, doi:10.1016/50278-4343(00)0012-7. Gordon, J.D.M., Bergstad, O.A., Figueiredo, I., Menezes, G., 2003. Deep-water fisheries of the Northeast Atlantic. I. Description and current trends. J. Northw. Atl. Fish. Sci. 31, 137–150. Haedrich, R.L., Barnes, S.M., 1997. Changes over time of the size structure in an exploited shelf fish community. Fish. Res. 31, 229–239. Hilborn, R., Branch, T.A., Ernst, B., Magnusson, A., Minte-Vera, C.V., Scheuerell, M.D., Valero, J.L., 2003. State of the World’s Fisheries. An. Rev. Environ. Resour. 28, 359–399. ICES, 2008. Report of the Working Group on the Biology and Assessment of DeepSea Fisheries Resources (WGDEEP), 3–10 March 2008. ICES CM 2008/ ACOM:14. ICES, Copenhagen. 531pp. Jones, E., Beare, D., Dobby, H., Trinkler, N., Burns, F., Peach, K., Blasdale, T., 2005. The potential impact of commercial fishing activity on the ecology of deepwater chondrichthyans from the west coast of Scotland. ICES CM N16, 1–19. Koslow, J.A., 1996. Energetic and life-history patterns of deep-sea benthic, benthopelagic and seamount-associated fish. J. Fish Biol. 49 (Suppl. A), 54–74. Koslow, J.A., Boehlert, G.W., Gordon, J.D.M., Haedrich, R.L., Lorance, P., Parin, N., 2000. Continental slope and deep-sea fisheries: implications for a fragile ecosystem. ICES J. Mar. Sci. 57, 548–557, doi:10.1006/jmsc.2000.0722. Lorance, P., Bergstad, O.A., Large, P.A., Gordon, J.D.M., 2008. Grenadiers in the Northeast Atlantic-distribution, biology, fisheries and their impacts, and developments in stock assessment and management. Am. Fish. Soc. Symp. 63, 365–397. Lorance, P., Dupouy, H., 2001. CPUE abundance indices of the main target species of the French deep-water fishery in ICES Sub-areas V–VII. Fish. Res. 51, 137–149, doi:10.1016/S0165-7836(01)00241-7. Minto, C., Nolan, C.P., 2006. Fecundity and maturity of orange roughy (Hoplostethus atlanticus, Collett 1889) on the Porcupine Bank, Northeast Atlantic. Env. Biol. Fishes 77, 39–50, doi:10.1007/s10641-006-9053-0. Moranta, J., Massutı, E., Palmer, M., Gordon, J.D.M., 2007. Geographic and bathymetric trends in abundance, biomass and body size of four grenadier fishes along the Iberian coast in the western Mediterranean. Prog. Oceanogr. 72, 63–83, doi:10.1016/j.pocean.2006.09.003. Morato, T., Watson, R., Pitcher, T.J., Pauly, D., 2006. Fishing down the deep. Fish Fish. 7, 24–34, doi:10.1111/j.1467-2979.2006.00205.x. Orlov, A.M., Iwamoto, T., 2009. Grenadiers of the World Oceans: Biology, Stock Assessment, and Fisheries. American Fisheries Society, USA 484pp. Pauly, D., Christensen, V., Dalsgaard, J., Froese, R., Torres, F., 1998. Fishing down marine food webs. Science 279, 860–863, doi:10.1126/science.279.5352.860. Roberts, C.M., 2002. Deep impact: the rising toll of fishing in the deep sea. Trends Ecol. Evol. 17, 242–245, doi:10.1016/S0169-5347(02)02492-8. Schnute, J.T., Richards, L.J., 2002. Surplus production models. In: Hart, P.J.B., Reynolds, J.D. (Eds.), The Handbook of Fish Biology and Fisheries. Blackwell Scientific Publishing, Cambridge, pp. 105–126.
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Stable abundance, but changing size structure in grenadier fishes (Macrouridae) over a decade (1998–2008) in which deepwater fisheries became regulated Francis Neat , Finlay Burns Marine Scotland-Science, Marine Laboratory, P.O. Box 101, 375 Victoria Road, Aberdeen AB11 9DB, UK
a r t i c l e in f o
a b s t r a c t
Article history: Received 7 October 2009 Received in revised form 17 December 2009 Accepted 27 December 2009 Available online 7 January 2010
A ten-year time series (1998–2008) from a trawl survey of the continental slope of the NE Atlantic was analyzed to assess temporal variation in the abundance and length frequency of seven species of deepwater grenadier fish. This period coincided (in 2003) with the regulation of deepwater fisheries in this area. None of the species declined in numbers or biomass over the period, and 2 species significantly increased. This suggests that the declines in abundance of these deepwater species following the onset of fishing in the 1970s may now have stabilized, albeit at much lower levels than the virgin biomass. Although two metrics of body size (mean length and maximum length) did not show any evidence for consistent decrease over time, there were significant changes in the overall length–frequency distributions. The species found in shallower depths (500 m) had a greater number of larger individuals in 2008 whereas those found deeper (1500 m) tended to have a greater number of smaller individuals. This suggests the presence of a lagged indirect effect of fishing on species that live beyond the actual depths that fishing takes place. & 2010 Elsevier Ltd. All rights reserved.
Keywords: Deepwater fisheries Coryphaenoides CPUE Length frequency Time series
1. Introduction The continental slopes of the world’s oceans separate the shelf seas from the abyssal plains and are characterized by steep gradients in seabed conditions and enhanced productivity (Gage and Tyler, 1993). Although slope environments represent only a small fraction of the area of the oceans, they support large and diverse fish assemblages. Consequently they are the focus of the majority of deepwater trawl fisheries. Trawlers began fishing the slopes of the NE Atlantic in the1970s where they operated as deep as 1500 m, although most effort was at depths around 1000 m where commercially valuable species such as roundnose grenadier (Coryphaenoides rupestris), orange roughy (Hoplostethus atlanticus), blue ling (Molva dipterygia), black scabbard (Aphanopus carbo) and deepwater sharks (e.g. Centrophorus squamosus) were most abundant (Gordon, 2001; Gordon et al., 2003). The fish assemblage in this area was also well studied in the 1970s just prior to the establishment of the major commercial deepwater trawl fisheries (Gordon, 1986). As research progressed it became apparent that many deepwater fish attained very old ages, had slow growth rates and were very late to mature, the combination of which made them unproductive and highly vulnerable to
Corresponding author. Tel.: + 44 1224 295516; fax: + 44 1224 295511.
E-mail address: [email protected] (F. Neat). 0967-0637/$ - see front matter & 2010 Elsevier Ltd. All rights reserved. doi:10.1016/j.dsr.2009.12.003
over-exploitation (Bergstad, 1995; Koslow et al., 2000; Morato et al., 2006). The example often cited is that of orange roughy (Clark, 2001) which is among the most vulnerable of all deepwater fishes because of its slow growth, late age at maturity (Minto and Nolan, 2006), extreme longevity (Andrews et al., 2009) and aggregating behavior. Furthermore, since the by-catch taken by deepwater fisheries is often large (Allain et al., 2003) and suffers total mortality due to the barotrauma of being brought from great depths, it became apparent that the entire fish assemblage of the slope could be highly vulnerable. Such concerns seem to have been borne out, not only by the decline in catches (Lorance and Dupouy, 2001), but also by two recent studies based on fisheries-independent data. The first, an analysis of data from the western North Atlantic suggested dramatic declines (up to 90%) in abundance of several deepwater slope species since the onset of deepwater fishing (Devine et al., 2006). The second, an analysis of the numerical abundance of the deepwater fish from the slope of the Porcupine bank area in the NE Atlantic suggested that by the 1990s numbers had declined to approximately half of what was estimated prior to exploitation (Bailey et al., 2009). Declines in abundance are, however, inevitable population consequences of exploitation (Haedrich and Barnes, 1997; Pauly et al., 1998; Hilborn et al., 2003), and theoretically a biomass reduction of half may be required to attain maximum productivity (Schnute and Richards, 2002). The declines in deepwater species were
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certainly dramatic; however, given that the absolute biomass and population dynamics of the fishery were largely unknown, it is difficult to determine whether this is indicative of terminal depletion or just the expected trajectory of a fishery on its way to reach maximum productivity. If the former, it would be expected that the stocks would continue to decline to commercial extinction. On the other hand, it could be argued the perception of the severity of the decline is due to the fact that unlike most shallow-water fisheries that have faced centuries of exploitation, the declines in deepwater stocks can be referenced to data that were collected when the stock was in a recently unexploited, ‘virgin’ state. If so, and the decline reflects a particularly fast case of ‘the cropping of the stock’, then it would be predicted that the stocks eventually stabilize at a lower absolute biomass. These two alternative interpretations need not be mutually exclusive and may both underlie observed patterns in deepwater fisheries. The problem is teasing the two apart and this has been hampered until now by a lack of fisheries-independent information from the years that followed the initial intensive exploitation. As the deepwater fisheries grew unregulated, so did concerns over their sustainability. In 2003, a total allowable catch (TAC) system was implemented for managing the ten major deepwater stocks in the NE Atlantic. This effectively capped any further expansion of deepwater fisheries until such time that there was good evidence that any increase could be sustainable. The International Council for the Exploration of Sea (ICES) held the view that most stocks were ‘probably outside safe biological limits’ and advised on precautionary fishing effort reductions. Since 2005 TACs and subsequently fishing effort and landings of most deepwater species have declined (ICES, 2008). For example, the TAC for roundnose grenadier decreased from over 5000 tonnes to under 4000 tonnes with landings declining to under 3000 tonnes in 2008 (Lorance et al., 2008; ICES, 2008). The consequences of such regulations have had for the populations of slope-dwelling fish remained largely unknown. There is therefore a pressing need for information from the period following the initial exploitation and over the period during which the fisheries were regulated, i.e. the past decade. Since 1998 Marine Scotland (formerly Fisheries Research Services) has undertaken a deepwater trawl survey of the continental slope to the west of Scotland, the data of which can be used to address these issues. Here we estimate recent trends in the abundance and size structure of one family of slope-dwelling fishes, the grenadiers (family: Macrouridae). Only the roundnose grenadier is commercially targeted as part of a mixed trawl fishery, but the six other species may be taken as by-catch. Grenadiers were chosen as the focus for several reasons: (1) they are numerous enough to obtain good estimates of abundance and length frequencies, (2) The seven species commonly found span the full extent of the depth gradient of the continental slope in this region (Lorance et al., 2008). Therefore some species are within depth strata directly exposed to fishing while others are beyond the depth of current fishing limits ( 1200 m), (3) they are one of the better studied families of deep-sea fishes (Moranta et al., 2007; Orlov and Iwamoto, 2009) and so information on life-histories is available to aid interpretation of results. Overall therefore, this family of fishes is a good model group for investigating fishing impacts on the continental slope. We first address the question of whether there is any evidence of a change over the past decade in the abundance of the grenadiers in terms of numbers and biomass. Secondly because exploitation often has consequences for the size structure of populations, we ask if there is any evidence of a change over the past decade in length– frequency distribution and body size.
435
2. Materials and methods 2.1. The Marine Scotland deepwater survey Marine Scotland used FRV Scotia to undertake a deepwater bottom trawl survey of the continental slope in the years 1998, 2000, 2002, 2004, 2005, 2006, 2007, 2008. The survey took place in September of each year and trawling was carried out mainly during daylight hours with a typical duration of 2 h bottom time. The full survey area (located within ICES area VIa) stretched along and down the slope from 551 to 591N at approximately 91W and covered depths from 300 to 1900 m. This full dataset was used to investigate the depth distribution patterns of the different Macrourid species. However, not all of this area was consistently surveyed at all depth strata and all years. Therefore for the purposes of modeling abundance and length frequency, data were derived from hauls taken from an area that spanned 56–58.51N at approximately 91W and at depths of between 400 and 1600 m (Fig. 1). This area covers five ICES statistical rectangles (41EO, 42EO, 43EO, 44EO, 45EO). A total of 110 hauls were available for analysis which gave near complete coverage in terms of ICES squares, depth strata (500, 1000, 1500 m) and years, the only exception being 1500 m in the year 1998 (Table 1). Trawling depth was kept to 7100 m of each depth stratum, but typically this did not vary more than 50 m within any particular haul.
2.2. Trawl gear specifications A bottom trawl (BT184) was rigged with 2100 rock-hopper ground gear (Jackson Trawls Ltd, Peterhead, UK), 1700 kg doors of area 5.82 m2 (Morgere, St.Malo, France), 100 m sweeps and floats rated to 2500 m. Warp to bottom-depth ratio ranged between
Fig. 1. A map of the survey area showing depth shading with 500, 1000 and 1500 m depth contours (white), the ICES statistical rectangle numbers and trawling positions (black lines). The island of St. Kilda (UK) is shown for reference.
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3. Results
Table 1 Numbers of hauls available per depth category per year for modeling. Year
500 m
1000 m
1500 m
Total
1998 2000 2002 2004 2005 2006 2007 2008 Total
5 6 4 6 4 5 5 5 40
5 5 4 4 4 5 5 5 38
NA 5 5 4 4 5 5 5 33
10 16 13 14 12 15 15 15 110
approximately 2.5:1 and 2:1 decreasing gradually with depth. The cod-end was fitted with an internal liner with 20 mm mesh size which ensured retention of fishes as small as a few cm in length. The entire catch was sorted and identified to species level. For each haul, the total weight for each species was recorded. Due to the tendency for the fragile tails of Macrourid species to break off in the process of being trawled, total length was not a suitable measure; instead length was measured as the distance (cm) from the snout to the first ray of the anal fin (the pre-anal fin length—PAFL). The total number of individuals for each species was determined by measuring the PAFL for each individual fish and creating a length–frequency distribution. There were instances where a species was too numerous for all fish to be measured and a subsample by weight was taken with the final numbers per cm size class raised accordingly. Efforts were made to ensure this subsample was representative of the entire catch, i.e. baskets were sampled randomly with respect to the sequence by which they were sorted and weighed. 2.3. Data and statistical analysis Catch per unit effort (CPUE) data were analyzed using generalized linear models (GLM). Two measures of CPUE were considered: numbers per hour (N h 1) of trawling and kilograms per hour (kg h 1) of trawling. To account for the over-dispersed nature of the N h 1 data the GLM was fitted using a negative binomial distribution. For the kg h 1 data the GLM was fitted with a gamma distribution. Although the time series spanned 10 years, this was of insufficient length to formally test for autocorrelation within the data. Instead the residuals of the models were inspected with respect to year to see if there were any indications of cyclic temporal trends. Similarly, to assess the degree of spatial autocorrelation the model residuals were examined with respect to sequential ICES statistical rectangles. Year, depth category and ICES squares (effectively latitude) were treated as factors in the models, and if interactions between factors were not significant the interaction terms were removed. In most cases there was one haul for each year, depth and ICES square, although some missing values were inevitable due to invalid hauls (e.g. net getting stuck and/or being damaged on the seabed) or constraints (e.g. weather) on completing all hauls in some years. For each species only hauls from their core depth distribution (see Section 3.1) were used. For analysis of mean and maximum length (per haul) a linear regression was used to model mean length or maximum length from each haul. The Kolmogorov–Smirnoff (K–S) test was used to determine if the length–frequency distribution of the first year of the time series was significantly different from the last. Any consistent trends in the direction of distribution shift over years were taken to be indicative of a temporal pattern (a shift to left indicates a greater proportion of smaller fish and vice versa) whereas a lack of consistency over years was taken to represent stochastic variability.
3.1. Species depth distributions and assignment of core depth strata Although 10 species of grenadier have been recorded on the survey, only 7 species are encountered in sufficient frequency to model temporal and spatial patterns effectively. These were Caelorhyncus caelorhyncus, Nezumia aequalis, C. rupestris, Coryphaenoides guentheri, Coryphaenoides mediterraneus, Caelorhyncus labiatus and Trachyrinchus murrayi. There were clear patterns of abundance (square root N h 1) with respect to depth (Fig. 2). The distribution of C. caelorhyncus ranged from o300–1000 m, with 82% of individuals caught between 400 and 600 m. Its core depth stratum was taken to be 500 m. N. aequalis was found deeper between 400 and 1300 m, with 95% of individuals caught between 700 and 1100 m. Its core depth stratum was taken to be 1000 m. C. rupestris had the widest depth distribution ranging from 500 to 41800 m with 91% of individuals caught between 900 and 1600 m. To account for this both 1000 and 1500 m were taken to reflect its core distribution. C. guentheri was found between 800 and 1800 m with 98% of individuals caught between 1300 and 1800 m. Its core depth stratum was taken to be 1500 m. C. mediterraneus was found between 1200 and 1800 m with peak numbers occurring between 1300 and 1800 m. Its core depth stratum was taken to be 1500 m. C. labitaus was found between 800 and 1800 m with 99% of individuals caught between 1300 and 1800 m. Its core depth stratum was taken to be 1500 m. T. murrayi was found between 800 and 1600 m with 91% of individuals caught between 1200 and 1600 m. Its core depth stratum was taken to be 1500 m. These observations are highly comparable to data presented by Lorance et al. (2008; Table 1) for data collected from the same area in the 1970s and 1980s suggesting the bathymetric distributions of the species are stable over decadal time spans. 3.2. Variation in relative abundance, mean size and length frequency For C. caelorynchus and N. aequalis data was available for all years in the time series, but the other species had core depth ranges that extend deeper than 1000 m, data was available for all years except 1998. The statistical results of the GLM and linear model analyses are presented in Table 2. 3.2.1. C. caelorynchus (500 m) Both numbers per hour (N h 1) and kilograms per hour (kg h 1) increased significantly between 1998 and 2008 (Fig. 3a). There was no effect of ICES square and no significant interaction. There was no trend over time with respect to mean length, although there was significant variability between ICES squares. Maximum length however did show a significant increase over years (Fig. 3b). The length–frequency distribution in 1998 was significantly different from that in 2008 (K–S test, D=0.64, P o0.001) with a higher frequency of larger individuals in 2008. This pattern also held for 2000 and 2002 suggesting a directional trend from the early to later years (Fig. 3c). 3.2.2. N. aequalis (1000 m) Both N h 1 and kg h 1 showed a trend for increasing numbers in later years, although this was only significant for kg h 1 reflecting a sudden increase in 2004 that remained high thereafter (Fig. 4a). There was also a significant effect of ICES square with greater numbers and biomass in more southerly squares. There was no significant trend over time with respect to mean or maximum length (Fig. 4b) and no significant variability between ICES squares. The length–frequency distribution of this species in
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437
70
C. caelorinchus
Sqrt Abundance (N.h-1)
60
N. aequalis 50
C. rupestris 40
C. guentheri 30
C. mediterraneus 20
C. labiatus 10
T. murrayi 0 300 400 500 600 700 800 900 1000 1100 1300 1500 1700 1800 Depth (m) Fig. 2. Bathyal distribution of the 7 species of grenadier expressed as square root abundance and standardized to 1 h. of trawling. Note no hauls were taken at depths less than 300 m.
Table 2 Results of GLM analyses of CPUE and size metrics for each species. Species depth (m) number of hauls
Variable
Year and trend
ICES Square
Depth
C. caelorynchus 500 m n = 38
N h1 kg h 1 Mean length Max length
Po 0.001 (m) Po 0.001 (m) NS, P= 0.11 (–) Po 0.01 (m)
NS, P =0.11 NS, P =0.42 P o 0.01 P o0.01
N/A N/A N/A N/A
N. aequalis 1000 m n = 36
N h1 kg h 1 Mean length Max length
NS, P= 0.07 (m) P= 0.01 (m) NS, P= 0.47 (–) NS, P= 0.57 (–)
P o0.001 P =0.015 NS, P =0.43 NS, P =0.39
N/A N/A N/A N/A
C. rupestris 1000–1500 m n = 65
N h1 kg h 1 Mean length Max length
NS, P= 0.32 (–) NS, P= 0.13 (–) NS, P= 0.89 (–) Po 0.001 (m)
P o0.05 P =0.012 P =0.012 P o 0.01
P o0.001 P o0.001 P o0.001 P o0.05
C. guentheri 1500 m n = 33
N h1 kg h 1 Mean length Max length
NS, NS, NS, NS,
(–) (–) (–) (–)
P =0.001 NS, P =0.39 P =0.002 NS, P =0.09
N/A N/A N/A N/A
C. mediterraneus 1500 m n = 33
N h1 kg h 1 Mean length Max length
Po 0.001, (m) NS, P= 0.08 (m) NS, P= 0.12 NS, Po 0.17
P =0.001 P =0.015 NS, P =0.16 NS, P =0.49
N/A N/A N/A N/A
C. labiatus 1500 m n = 33
N h1 kg h 1 Mean length Max length
Po 0.001 (m) NS, P= 0.44 (–) Po 0.001 (k) P= 0.03 (-)
P o0.001 P =0.019 P =0.72 NS, P =0.78
N/A N/A N/A N/A
T. murrayi 1500 m n = 33
N h1 kg h 1 Mean length Max length
NS, NS, NS, NS,
P o0.001 P =0.019 NS, P =0.69 NS, P =0.69
N/A N/A N/A N/A
P= 0.56 P= 0.17 P= 0.21 P= 0.33
P= 0.28 (–) P= 0.41 (–) P= 0.31 (–) Po 0.39 (–)
N/A = not applicable, P =P-value, NS =non-significant, (m) = increasing trend, (–) = no trend, (k) = decreasing trend.
1998 was significantly different from that in 2008 (K–S test, D= 0.23, Po0.001; Fig. 4c) with larger individuals more frequent in 2008, however, this pattern was not consistent across other years and may reflect short-term stochastic variation as opposed to a longer term trend.
3.2.3. C. rupestris (1000–1500 m) For both N h 1 and kg h 1 (Fig. 5a) there was no effect of year. There was a significant effect of depth with greater numbers and biomass at 1500 m and there was significant variability between ICES square, although these differences were not consistently
related to latitude. There appears to have been no significant change in the mean length of this species over years. Mean and maximum length was significantly greater at 1500 m and varied significantly between ICES squares. The maximum length of this species showed a small but significant increase over the time period (Fig. 5b), although the presence of a significant interaction between depth and year suggested the increase was apparent only at 1500 m. The cumulative length–frequency distribution of this species in 2000 was significantly different from that in 2008 (K–S test, D= 0.42, Po0.001; Fig. 5c). This appears to be related mainly to a decrease in the number of small individuals ( o9 cm) in later years.
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Figs. 3–9. Indices of abundance, size metrics and length–frequency data for each species in each year: C. caelorhyncus (Fig. 3), N. aequalis (Fig. 4), C. rupestris (Fig. 5), C. guentheri (Fig. 6), C. mediterranaeus (Fig. 7), C. labiatus (Fig. 8) and T. murrayi (Fig. 9). For each species: (a) Mean + - 95% cls CPUE data (N h 1—filled circles, solid line and kg h 1open circles, dashed line); (b) mean PAFL (light gray bars, plus standard deviation) and average maximum PAFL (dark gray bars, plus upper range) and (c) cumulative percentage of individuals for each length class. Solid black curve = 1998 or 2000, dashed curve =2008, gray curves =intervening years.
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3.2.4. C. guentheri (1500 m) For neither N h 1 nor kg h 1 (Fig. 6a) was there any significant effect of year. There were however significant effects of ICES square with numbers steadily increasing from south to north. There was no significant trend over time with respect to mean or maximum length (Fig. 6b), although there was significant variability between ICES squares. The cumulative frequency distribution of this species in 2000 was significantly different from that in 2008 (K–S test, D= 0.08, Po0.05; Fig. 6c) with higher frequencies of smaller individuals in 2008. 3.2.5. C. mediterraneus (1500 m) For both N h 1 and kg h 1 (Fig. 7a) there was an increasing trend over time, although this was only significant for N h 1. There was a significant effect of ICES square (P= 0.001) with numbers and biomass steadily increasing from south to north. There was no significant trend over time with respect to mean or maximum length (Fig. 7b) and no significant variability between ICES squares. The cumulative frequency distribution of this species in 2000 was significantly different from that in 2008 (K– S test, D=0.27, Po0.001; Fig. 7c) with smaller individuals more frequent in 2008. This pattern was consistent in each subsequent year. 3.2.6. C. labiatus (1500 m) For N h 1 (Fig. 8a) there was a significant effect of year with numbers increasing until 2008, which showed a downturn. There also was a significant effect of ICES square with numbers steadily increasing from south to north. For kg h 1 (Fig. 8a) there was no significant effect of year, but there was a greater biomass in more northerly squares. There was a significant effect of year with respect to mean length (Fig. 8b) with a consistent decline in mean size since 2002 and no significant variability between ICES squares. There was also a significant effect of year with respect to maximum size, although no consistent trend over years was apparent. The cumulative frequency distribution of this species in 2000 was significantly different from that in 2008 (K–S test, D= 0.30, Po 0.001; Fig. 8c) with smaller individuals more frequent in 2008. The pattern was not however consistent from year to year. 3.2.7. T. murrayi (1500 m) For neither N h 1 nor kg h 1 (Fig. 9a) was there any effect of year. There was an effect of ICES square with numbers and biomass steadily increasing from north to south. There was no significant trend over time with respect to mean or maximum length (Fig. 9b) and no significant variability between ICES squares. The cumulative frequency distribution of this species in 2000 was not significantly different from that in 2008 (K–S test, D= 0.25, NS; Fig. 9c) and showed little difference across all years.
4. Discussion Most fish stocks will show an initial downturn in abundance following exploitation (Schnute and Richards, 2002); this is not unique to deepwater fisheries. What is unique to deepwater fisheries is that only 40 years ago they were in a virgin biological state, and that they underwent a particularly fast decline immediately following the onset of fishing (Devine et al., 2006; Bailey et al., 2009). Deepwater fish differ from most shallowwater fish in being very long-lived, slow growing and late maturing (Koslow et al., 2000). Consequently they are less productive and likely to be highly vulnerable to over-exploitation (Morato et al., 2006). Much concern and debate has been raised
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over stock declines and whether they can be sustainably exploited (Roberts, 2002). Some deepwater species such as orange roughy (Clark, 2001) almost certainly cannot be fished sustainably because catches would have to be so low as to be unprofitable. For other deepwater species, however, biological and economic factors may be rather different. Provided fisheries for such species are carefully regulated on the basis of good data there may be sustainable levels of profitable exploitation. The results from this study demonstrate that the grenadiers from the NE Atlantic have not declined in the past ten years. Although the current area studied is some 200 miles north of the study area of Bailey et al. (2009), the slope-dwelling fishes would most likely have undergone a similar decline because the main commercial trawl fisheries extended along most of the slope (Gordon et al., 2003). Assuming this to be the case, it would appear that the decline may have now stabilized. The stocks, however, show little sign of recovering, and in absolute terms they are probably still a fraction of what they were at of preexploitation virgin biomass levels. This is perhaps not surprising given that estimates of stock recovery time for C. rupestris range from 14 to 80 years (Baker et al., 2009). Thus while the rate of population decline caused by fishing appears to have been particularly fast in these deepwater species the fact that they do appear to be stabilizing suggests they are not in terminal decline. Unfortunately it is difficult to compare directly our abundance estimates with those collected in the 1970s because of the differences between survey gear, research vessels and area. Nevertheless, as indices of relative abundance, the results are clear-cut and because the survey design and analyses controlled for or accounted for the effects of season, depth and sampling area, we are confident they reflect genuine temporal patterns. Furthermore it is unlikely that the present results have arisen from a shift in core depth distribution of each species over the time period in part because the depth distributions for these species reported here are highly comparable to these reported in the same area in the 1970s (Lorance et al., 2008). The relative stability since 1998 may reflect the introduction (in 2003) of a management regime including TACs for commercial deepwater species which included roundnose grenadier. Since 2005 TACs have generally declined and some directed fisheries, such as for orange roughy and deepwater sharks are being phased out altogether. Thus fishing pressure has been alleviated and these data cautiously suggest that it may have been sufficient to prevent the further decline of the grenadier fishes. It is interesting that those species that showed an increase in abundance were the shallower, smaller species such as C. caelorynchus and N. aequalis. This may be because of their lower age and size at maturity (Coggan et al., 2000) compared to species such as C. rupestris (Bergstad, 1990). Alternatively because these smaller species occur in the depth strata at which the most intensive trawling activity took place, they may have experienced a greater relative reduction in fishing pressure and the trend is thus more pronounced. Although the grenadiers are a very diverse family (Orlov and Iwamoto, 2009) with a broad range in body sizes, growth rates and longevity, their energetics and life-history are fundamentally different from the most vulnerable deepwater species that include seamount-associated species, such as orange roughy (Koslow, 1996) and deepwater sharks which have very ˜ o´n et al., 2006). Consequently the low reproductive rates (Ban present findings should not be taken as indicative that all deepwater species may be expected to show signs of abundance stabilization following the regulation of fisheries. With respect to size-based metrics there were no signs that the mean or maximum size of individuals had decreased in the past decade. These two metrics are often sensitive indicators of fish populations in process of being over-exploited (Haedrich and
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Barnes, 1997; Bianchi et al., 2000). Ten years may be insufficient time to observe such changes in long-lived species and it is perhaps more informative to look at the overall size structure of the populations. Here there were some interesting patterns over time particularly when all the species are compared across the full depth range of the slope. The one species (C. caelorynchus) with a core depth distribution on the upper slope at 500 m had a greater number of larger individuals in the population in 2008 than in 1998, 2000 and 2002, suggesting a directional trend. This trend was not evident in those species found at 1000 m (N. aequalis and C. rupestris) and was in the opposite direction in 3 out the 4 species found at 1500 m (C. guentheri, C. mediterraneus and C. labiatus). The increase in the number of larger individuals in the upper slope species may reflect the recent reduction in fishing mortality with individuals living longer and growing larger. On the lower slope, the increase in the number of smaller individuals into the populations may be related to what Bailey et al. (2009) describe as the impact of fishing extending well beyond the actual depth of fishing. Although Bailey et al. (2009) noticed such effects with respect to abundance, it is likely the size structure of the populations will also be affected. The mechanisms underlying such changes are not clear, but one possibility is an indirect effect that arises from the reduction of apex predators such as deepwater sharks (Jones et al., 2005) and blue ling (Lorance and Dupouy, 2001), which in turn has lead to an increase in the survival probability of small individuals.
5. Conclusion The data presented here suggest that the alarming declines in deepwater species reported after the fisheries began in the Rockall Trough may have stabilized in at least one major group of slope-dwelling fishes. It is cause for cautious optimism that the recent reduction and regulation of deepwater fishing in this area may have been sufficient to prevent further decline. There is, however, little sign of recovery to former levels and without more information on stock productivity (such as growth rates) it would be premature to comment on whether fishing for species such as roundnose grenadier could be considered sustainable.
Acknowledgments The research was supported by the Scottish Government (MF0763). We are especially grateful to FRV Scotia Fishing Masters Neville Ball and Peter Carmichael, their crew and the many Marine Laboratory staff that helped collect these data, notably Kevin Peach and Emma Jones. We further thank Peter Wright, Neil Campbell and Bill Turrell for comments on the paper, and M. Bacon, I. Priede and three anonymous referees for their critical and constructive review of the paper. References Allain, V., Biseau, A., Kergoat, B., 2003. Preliminary estimates of French deepwater fishery discards in the Northeast Atlantic Ocean. Fish. Res. 60, 185–192, doi:10.1016/S0165-7836(02)00031-0. Andrews, A., Tracey, D.M., Dunn, M.R., 2009. Lead-radium dating of orange roughy (Hoplostethus atlanticus): validation of a centenarian life span. Can. J. Fish Aquat. Sci. 66, 1130–1140, doi:10.1139/F09-059. Bailey, D.M., Collins, M.A., Gordon, J.D.M., Zuur, A.F., Priede, I.G., 2009. Long-term changes in deep-water fish populations in the northeast Atlantic: a deeper
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