In the early months of the year, mature haddock migrate southwards from the Barents Sea to the west coast of Norway where they spawn, mostly along the shelf edge, March–May. After spawning, the adult fish migrate northwards to their summer feeding grounds in the Barents Sea. The extent of the summer–winter haddock distribution within the Barents Sea varies from year to year in response to temperature and the distribution and abundance of prey species. Their migrations extend to the Svalbard Shelf in the north and eastwards towards Nova Zemlya but the extent of the distribution is limited by the Polar Front associated with the ice sheet. During the spawning migration of capelin, haddock prey on capelin and their eggs on the spawning grounds. When the capelin abundance is low or when their areas do not overlap, haddock can compensate for the lack of capelin with other fish species such as young herring, or with euphausids and benthos, which are predominant in the haddock diet throughout the year. Density-dependent growth has been observed for this stock and the present growth rate is low. Cod is the main predator on haddock and this predation is included in the natural mortality used in the assessment. The predation by cod on haddock has been high in recent years due to the large cod stock size (ACOM NEAHad , 2012). 52 Variation in the recruitment of haddock has been associated with changes in the influx of Atlantic waters to the Barents Sea. Water temperature in the first and second years of the haddock life cycle is one of the factors that determine year-class strength; the probability of good recruitment is very low when the temperature is low. Additionally, a steep rise or fall of the water temperature shows a marked effect on the abundance of year classes. This information on environmental influence is not yet taken into account in the assessment (ACOM NEAHad , 2012).
A large proportion of the species targeted in the ECIFFF occupy top positions in the marine food web and are often classified as tertiary consumers including the majority of the shark species (Cortés, 1999), a number of the key teleost species (Ceccarelli & Ayling, 2010) and occasional ray species (Jacobsen & Bennett, 2013). Given this, one of more likely risks associated with this fishery will be the removal of top order predators and other (secondary consumers) from the food chain (Appendix 1). At a whole of fishery level, this risk will be due to their retention for commercial sale (harvesting) and the discarding of animals in a dead or moribund state due to their low value or where regulations prohibit their retention. For one of the more prominent species in this group, sharks, this risk will be highly relevant to non-S fishery symbol holders who are restricted through in-possession limits. The management regime for the ECIFFF includes a range of measures to protect larger individuals with higher reproductive potential e.g. minimum legal size limits, mesh size restrictions. These management initiatives are more developed for key species and minimum and maximum legal size limits often account for the size at sexual maturity. These measures increase the likelihood that an individual will reproduce at least once before they are harvested and will help to maintain long-term recruitment rates. These measures though will be less effective in the net fishery where there is increased potential for operators to interact with cohorts outside the prescribed size limits. Similarly, maximum legal size limits for sharks will be less influential as they do not apply to net operators with an ‘S’ fishery symbol. This is largely done to minimise the level of waste/discards in the fishery. The effective management of stocks will depend on a range of factors and the capture of individuals above or below the prescribed size limits does not automatically equate to a species being unsustainably fished. However, in the context of this (and future) ERAs this issue does present a risk as it has the potential to undermine the effectiveness of key management arrangements and or have an effect on recruitment rates for some species.
In terms of the biology and ecology of these species, the key factors to understand from the fisheries management point of view is i) that these species are all fairly fast-growing and short-lived, maturing young and reproducing in large quantities; ii) that they are highly mobile, and dependent on food resources (plankton etc.) which are patchy in time and space; and iii) that the distribution of their eggs, larvae and food are driven by oceanographic and weather conditions which vary from season to season and year to year. All these factors conspire such that their biomass (the total amount of fish available to the fishery at any given time), recruitment (juvenile fish entering the fishery) and geographic distribution can be extremely variable from year to year, and this variability is driven not so much by fishing as by weather and oceanographic patterns, making it very unpredictable. From the point of view of management of these fisheries, this variability has a good side and a bad side: on the good side, the high inherent growth rate of the population means that the population can be fished quite hard without having a long-term impact, and also that the population recovers rapidly from depletion (they are ‘resilient’); on the bad side, it is extremely difficult to set reference points, sustainable total catch limits etc. working from a baseline which is naturally extremely variable. This means that it may be more appropriate in some cases to have an adaptive, ‘rule of thumb’ type approach to management, rather than a rigid scientific structure for taking
(Cetorhinus maximus). Additional occasional (irregular vagrant) captures in Icelandic waters include, blue shark (Prionace glauca), thresher shark (Alophias vulpes), porbeagle shark (Lamna nasus) 101 and blue fin tuna (Thunnus thynnus) 102 . The porbeagle, thresher and blue shark, and the bluefin tuna are all rare vagrants as southern Icelandic waters are at on the northern boundary of their distribution 103 and their presence in Icelandic waters is dependent on the vagaries of the northern boundary to the warm water of the North Atlantic drift. All are pelagic feeders and are unlikely ever to be taken in demersal trawls, Danish seines or bottom- set gillnets. In contrast, the basking shark is vulnerable to capture in demersal fishing gear, both towed (personal observation) and static gear. Of the chimeras, it is the common rat-tail or rabbit fish (Chimaera monstrosa) that is most vulnerable to capture in the mixed demersal fishery as it is not confined to deep waters like other chimeran species 104 . Even so, catch rates appear to be very low (≤1 t, 2008) with concomitant minimal effects from the mixed demersal fisheries.
This fisheryassessment report for the Western Zone (WZ) of the South Australian Abalone Fishery (SAAF; see Figure 1.1) updates previous fisheryassessment and status reports for Region A (Chick et al. 2006; 2008; 2009; Stobart et al. 2011; 2012b) and Region B (Chick and Mayfield 2006; Chick et al. 2007; Stobart et al. 2010; 2012a) of the WZ. The most recent report for blacklip abalone (Haliotis rubra), hereafter referred to as blacklip, was the WZ status report that followed amalgamation of the management of Regions A and B, effective from 1 January 2014 (Stobart et al. 2014a). This report provides an analysis of fishery-dependent (FD) and fishery-independent (FI) data for blacklip in the WZ from 1 January 1968 to 31 December 2014 and is part of the South Australian Research and Development Institute (SARDI) Aquatic Sciences’ ongoing assessment program for the blacklip fishery in the WZ. It also includes a formal analysis of the fishery’s performance and stock status based on the harvest strategy described in the Management Plan (PIRSA 2012), which determines the (1) risk that stocks in the high and medium spatial assessment units (SAUs) are overfished and (2) zonal stock status. In the discussion, we assess the current status of the blacklip stocks in the WZ comparing the harvest strategy and traditional weight- of-evidence assessments using the National Fishery Status Reporting Framework (NFSRF; Table 1.1, Flood et al. 2012; 2014) adopted by Primary Industries and Regions South Australia Fisheries and Aquaculture (hereafter referred to as PIRSA) as the framework of choice for classifying fish stocks.
Participants noted that the likelihood of capturing seabirds while targeting Spanish mackerel was low but is known to have occurred. The species involved are usually restricted to boobies and gannets. Species that have attracted concern in other jurisdictions such as albatross are not generally caught in the Spanish mackerel fishery because of their limited natural distribution. Participants also agreed that because they are constantly in attendance of fishing gear, they are often able to pull the line away should they see a bird diving for the bait. This essentially avoids any chance of hooking a bird or having one become entangled in the line.
Barramundi in the Southern GoC stock are taken commercially as part of the Gulf of Carpentaria Inshore Fin Fish Fishery (GOCIFFF), which extends from Slade Point near the tip of Cape York Peninsula to the Queensland/Northern Territory border. The GOCIFFF is a multi-species fishery that includes an inshore (N3 symbol) commercial net fishery that harvests inshore species such as barramundi and king threadfin, and an offshore (N9 symbol) commercial net fishery that targets offshore species such as shark and grey mackerel. The inshore N3 fishery uses set mesh nets (i.e., gill nets) in rivers, on foreshores and in more offshore waters out to seven nautical miles. See Roelofs (2003) and Ward (2003) for a detailed description of the GOCIFFF, including commercial fishing methods. The GOCIFFF is managed separately from the East Coast Inshore Fin Fish Fishery, with different management arrangements applying in each fishery. The GOCIFFF requires a Wildlife Trade Operation (WTO) for export approval and protected species accreditation under the Commonwealth’s Environment Protection and Biodiversity Conservation Act 1999, to demonstrate that the fishery is operating under national sustainability guidelines.
The flooding in late December 2010 and January 2011 caused barramundi, stocked in upstream impoundments (such as Awoonga Dam) and waterways of the CEC stock region, to escape and become available to the commercial fishery. Our attempt to incorporate the number of barramundi fingerlings released within the Central East Coast stock (Appendix B) into the current assessment model did not provide stable results. Instead, we use a smoothing technique to remove the bias in catch and catch rates in the years 2011 to 2015 (as an index of abundance) due to stocked fish. We replaced the catch in each of the years 2011, 2012, 2013, 2014 and 2015 by the moving average of the four preceding years, see the green curve in Figure 39.
proceeds to the next site along the survey route. This approach is an alternative to roving surveys, which are typically used to monitor small-scale commercial and subsistence fisheries. For the roving survey, a closed circuit route is laid out which traverse the entire fishery (Pollock et al. 1997). The observer interviews and/or counts fishers as he/she moves through the area (El-shaarawy & Piegorsch 2002). However, a few sources of bias likely to occur when applying this survey method to this subsistence fishery encouraged the application of an alternative method here. First, the probability of intercepting the observer is proportional to the length of the fisher’s fishing trips, and fishers fishing for only a few minutes are not likely to be intercepted by the observer (Pollock et al. 1997). In this fishery, spending time at a fishing site increased the probability of recording catch from fishing gears such as castnet. Such fishing methods usually do not last longer than 15-20 minutes and can thus be easily missed during a roving survey. Similarly, BRS were better at detecting boat-based activities and roving fishing activities, such as spearfishing. Usually such activities could be detected only during the small amount of time when boats left or arrived on the island, or fishers roved in one fishing site. Second, roving survey neglects interviewing times, which may be grossly erroneous, depending on the interview duration (Pollock et al. 1994). In contrast, BRS are an unbiased method of estimation of fishing effort, since they take into account the time spent by the observer interviewing fishers (Robson & Jones 1989, Jones et al. 1990). As an additional benefit, BRS also provided the chance to engage with people fishing or located around the fishing sites.
The length-frequency samples (Figures 13 and 15) show a large amount of year- to-year variation in the size distribution of fish, probably more variation than is actually present in the population. This variation has continued to take place even since 1999 when the sampling methodology was scientifically standardised by the Long Term Monitoring Program (see examples at the end of section 4). It appears that even though the number of schools sampled is large (E. Jebreen, personal communication), there are factors such as weather and surf conditions, and presence of seaweed in the water, that can influence the size distribution of fish taken by the fishery throughout a sampling period. Since 1999, the sampling period has generally been a week, and samples have been taken several times over the Fraser Island season. The sampling strategy is believed to accurately record the distribution of fish caught by the fishery during sampling periods.
Monitoring protocols were developed to assess the subsistence reef fish fishery in the Torres Strait (Objective 1). Protocols developed in this study, and information collected, could be used when monitoring other similar Indigenous fisheries in Australia. Creel surveys, usually used to monitor recreational fisheries, were employed. Semi-structured interviews were also used to collect additional information on subsistence fishing and on perceptions of community members about fisheries issues. The long-term employment of a local research counterpart, a common strategy to monitor reef fish fisheries, was precluded in this study by several socio-economic and cultural conditions of the communities. This is a point to carefully consider when developing monitoring programs in the Torres Strait, since it can determine their extent and success.
All technical reports have been submitted and approved by FFC (Federal Fisheries Council) for the Bonaerense anchovy fishery (Hansen et al., 2013 and 2014). In the period 1990-2013, the average yields for north stock of anchovy were lower than 22,000 t / year, representing a fraction of point estimates of available abundance. Accordingly, in this period the annual value of γ(M) varied between 0.735 and 0.845 (mean = 0.803, SD = 0.028), reflecting the distribution of catches in different periods. The most striking results are the low magnitudes of Fy (mean = 0.04), well below the rate M = 1.01 natural mortality, confirming that the stock is under-exploited (Hansen et al., 2014). The catch of 2014 was extremely low (not exceeds 12,500 t).
As the boundary of the C1 Fishery includes both the Queensland EC and the GoC the risk characterisation table (Table 4) is based at the whole of fishery level. As the dynamics of the C1 Fishery on the Queensland EC and the GoC are vastly different, this complicates the situation surrounding the preliminary risk assessments. For example, preliminary risk ratings based at a whole of fishery level will mask inter-regional differences including the likelihood that fishing activities in the GoC will present a lower risk for a number of the ecological components. To account for this variance, two regional Risk Characterisation tables have been provided in Appendix 2. These provide a more detailed view on the risks posed by the fishery in each region and the likelihood that one or more of the fishing activities will contribute to an undesirable event under the current management framework. Similarly, target species in the C1 Fishery have different profiles including the areas of operation. These differences were accounted for in the Risk Identification stage where the majority of risk attributed to the mud crab fishery. It is important to note though that the risk ratings assigned in Table 4 take into consideration a range of factors including the ability of operators to target one or both species, mechanisms to control catch or effort and ancillary factors such as the appeal of the species to other sectors. In the case of blue swimmer crabs, the risk ratings contained in Table 4 could be considered precautionary in nature. In the event that the species is classified as high-risk in
RFISH participation was subject to a greater dropout rate than in the NRIFS-SWRFS surveys, and the assumption had to be made that participants who dropped out caught as many fish on average as the participants who remained until the end of the survey. In practice, participants are likely to have dropped out not merely from the survey but also from fishing, and therefore to have caught fewer fish than the participants that remained. If this occurred, the RFISH results would be overestimates. The RFISH methodology was also less clear in its handling of catches that were shared between members of a fishing party, only one of whom would have been a diary respondent. We do not intend these points as criticism of the RFISH program, but an acknowledgement that methodology for such surveys was refined over time. We have followed the recommendation of Fisheries Queensland in regarding the NRIFS- SWRFS estimates as the best available, and the RFISH ones as potential overestimates. Fishing locations were specified precisely in the SWRFS survey, but in the NRIFS survey they were specified only in broad regions. Therefore it was easy to split the SWRFS results into the Subregions used in the stock assessment. The SWRFS results are listed in Table 12. Inner-shelf fishing locations were defined to be those in the GMRMPA Bioregions (see Figure 6) that were close to shore and not frequented by commercial fishers in the logbook data; these were locations beginning with “RE”, “RF” or “RHC”, plus the part of Bioregion “RD” that was in the Princess Charlotte Bay Subregion.
The generalist category recognises these species are less vulnerable due to their wide range of niches they associate with. This may provide refugia from fishery and climatic impacts which in turn could provide a source of recruitment. Generalists also include colonizing species or ‘R’ strategists referring to the species’ ability to quickly reproduce to fill disturbed areas or quickly repair following damage. The specialist category recognises that some species are more vulnerable because the niches they occupy are restricted in some way and/or the species have developed specialised behaviours/modifications to occupy particular niche habitats. These taxa are likely to be affected by limited
From the early 1990s, commercial fishers in Queensland were encouraged to fish in deep-water areas to reduce the pressure on more heavily fished shallow water areas, particularly those of the Great Barrier Reef. Initial entry policy to what was then to be developed as a multi-hook longline and dropline fishery, was generous with all Queensland holders of L1, L2 or L3 line fishery symbols on a primary commercial fishing boat entitled to apply for an L8 endorsement that enabled the use of multi-hook gear (>6 hooks) in water greater than 200 m. By 1999, 40 such L8 endorsements had been issued but in October 1999 a ‘freeze’ was placed on the granting of further L8 licenses. Subsequently, more stringent entry restrictions resulted in the reduction of the number of vessels carrying L8 endorsements to 5 today (January 2016). During this period, deep-water stocks continued to be accessed by commercial line fishers who did not have an L8 endorsement, as the usual L1, L2 and L3 entitlements allowed the use of up to six hooks per line. This limitation in the quantity of gear still enables economically viable catches of fish to be taken in waters greater than 200m and an examination of commercial and charter logbook harvests from remote locations (CFISH grids) throughout Queensland shows high levels of line fishing effort. In addition to the expansion of commercial and charter fishing effort, there is also evidence of expanding recreational interest in fishing deep-water areas (Sumpton et al. 2013a). Some of this effort is known to target pearl perch.
There are several uncertainties with respect to the use of CPUE as an indicator of relative abundance for finfish in the LCF. These include: (i) differences among individual licence holders in the way that effort is reported; and (ii) environmentally-mediated changes in the amount of habitat available for fish (i.e. size of the fishable area), particularly in the Coorong estuary, which may affect the catchability of some species. A further uncertainty relates to use of fisher days as the preferred unit of effort to calculate CPUE for finfish, because this unit of effort does not account for the number of gill nets that were deployed on each fishing day. Whilst a fishery- independent method for estimating the relative abundance of Pipi has been developed (Ward et al. 2010) and implemented since 2007/08, it is important to address uncertainty around CPUE for finfish, because it provides the only long-term (30 years) time series of relative abundance. In addition, the structure of the current fishery PIs and RPs for finfish provide limited scope for accurate assessment of stock status for key species. This is because: (i) RPs were determined based on total catch and CPUE data from a fixed, relatively short time period (i.e. from 1984/85 to 2001/02) and do not consider data from the past 12 years; and (ii) there is limited consideration of the influence of other factors which may affect fishery performance (e.g. habitat availability, Murray Mouth opening and market demands).