- Email: [email protected]
Effects of recreational clam harvesting on eelgrass (Zostera marina) and associated infaunal invertebrates: in situ manipulative experiments
Aquatic Botany 73 (2002) 63–74
Effects of recreational clam harvesting on eelgrass (Zostera marina) and associated infaunal invertebrates: in situ manipulative experiments Bruce L. Boese∗ US Environmental Protection Agency, Coastal Ecology Branch, Hatfield Marine Science Center, 2111 SE Marine Science Drive, Newport, OR 973656-5260, USA Received 12 March 2001; received in revised form 17 December 2001; accepted 17 December 2001
Abstract The effect of recreational clam harvesting on eelgrass (Zostera marina L.) was experimentally tested by raking or digging for clams in experimental 1 m2 plots located in a Yaquina Bay (Newport, OR) eelgrass meadow. After three monthly treatments, eelgrass measures of biomass, primary production (leaf elongation), and percent cover were compared between experimental and control (undisturbed) plots. Benthic macro (retained on 0.5 mm mesh sieve) and mega (retained on 3 mm sieve) infaunal samples were also taken to compare species number and abundances. Results indicated that clam raking did not appreciably impact any measured parameter. In contrast, clam digging reduced eelgrass cover, above-ground biomass and below-ground biomass in measurements made 1 month after the last of three monthly treatments. Although differences between control and treatment plots persisted 10 months after the last clam digging treatment, these differences were not statistically significant. Approximately 10% of the eelgrass of Yaquina Bay is subjected to recreational clamming and as this activity is generally less intense than that employed in this study, it is unlikely that recreational clamming has a major impact on eelgrass beds in the Yaquina estuary. This conclusion should be viewed with caution as multi-year disturbances were not investigated and there are differences in sediment characteristics and clam abundances between experimental sites and those sites that are intensively harvested by the public. Published by Elsevier Science B.V. Keywords: Zostera marina; Sediment; Disturbance; Clamming
1. Introduction Eelgrass (Zostera marina L.) beds, are an important habitat in Pacific northwest (PNW) estuaries. Eelgrass meadows serve as a nursery ground for juveniles of commercially ∗ Tel.: +1-541-867-5019; fax: +1-541-867-4049. E-mail address: [email protected] (B.L. Boese).
0304-3770/02/$ – see front matter. Published by Elsevier Science B.V. PII: S 0 3 0 4 - 3 7 7 0 ( 0 2 ) 0 0 0 0 4 - 9
64
B.L. Boese / Aquatic Botany 73 (2002) 63–74
important species such as herring and as a refuge for juvenile salmonids (Griffin, 1997; den Hartog, 1977; Simenstad and Wissar, 1985; Levings, 1990). Eelgrass meadows are significant sites of primary production and eelgrass shoots can be utilized directly for food by some water fowl such as the western black brant (Kentula, 1982; Griffin, 1997), and indirectly by many species via consumption of detritus (Thayer et al., 1975). Eelgrass roots stabilize the sediment (Thayer et al., 1975) and the presence of eelgrass dampens wave energy which may serve to reduce erosion and to enhance larval settlement (Orth, 1992). Because of these characteristics, species abundances in eelgrass patches are usually greater than in other estuarine habitats (Everett et al., 1995). Seagrasses may be impacted by anthropogenic physical disturbances such as dredging (Zieman and Zieman, 1989), prop scarring (Zieman, 1976; Matthews et al., 1991; Sargent et al., 1994), and boat moorings (Walker et al., 1989; Clinton et al., 1997) which directly remove biomass and alter the physical environment. Mechanical disturbances associated with commercial shellfish and bait harvesting operations have also been shown to adversely affect seagrasses (Fonseca et al., 1984; Peterson et al., 1987; Everett et al., 1995; Orth et al., 1998) and to reduce mudflat biodiversity (Brown and Wilson, 1997). Many of these studies suggested that the resulting reductions in seagrass biomass may take years to recover to predisturbance levels. One seagrass physical disturbance mechanism that has not been studied is recreational clamming. Recreational clamming is a significant activity in PNW estuaries which may have a direct impact on eelgrass and its associated biota. In Yaquina Bay, Newport, OR, gaper (Tresus capax) and butter clams (Saxidomus giganteus) are harvested by digging in lower intertidal Z. marina meadows while cockles (Clinocardium nuttalli) are often raked from the sediment surface of higher intertidal areas. Clamming activity is not limited to intertidal areas of eelgrass meadows as some clam diggers wade at low tide into 1 m of water to dig subtidal clams. Thus, this mechanical disturbance could affect eelgrass to −2 m mean lower low water (MLLW) in Oregon estuaries. As human coastal populations and associated tourism increases, the potential for physical disturbance from these recreational activities is also likely to increase. The objective of the present research was to examine the effects of one season of mimicked recreational clam digging and raking on eelgrass (Z. marina) and associated macro and mega infaunal species. 2. Methods 2.1. Site selection Yaquina Bay, the fifth largest estuary in Oregon, encompasses an area of approximately 1582 ha with 35% intertidal and 65% subtidal (Oregon Estuaries, 1973). Approximately 100 ha of this area is covered by Z. marina which is almost all intertidal (USEPA, Coastal Ecology Branch, unpublished data). Small patches of Zostera japonica are also present but are limited to a high intertidal fringe near the shore line (Fig. 1). The majority of the estuary’s eelgrass (≈85 ha) is located within the Yaquina Bay estuary’s embayment, which can be delineated into three large distinct intertidal regions: Sally’s Bend, Idaho Flat, and
B.L. Boese / Aquatic Botany 73 (2002) 63–74
65
Fig. 1. Intertidal Sally’s Bend showing extensive Z. marina meadow with some fringing Z. japonica patches located high in intertidal near shore line: (A) clam raking study site; (B) clam digging study site.
Raccoon Flat. Sally’s Bend was chosen as the site for this study, because it has the most extensive eelgrass meadows in Yaquina Bay, is relatively flat so that all experimental plots were within a narrow tide height range, and because public access is limited as the central areas of the bed are only accessible at low tide via a boat (Fig. 1). In order to further ensure that these areas would not be disturbed by recreational clam harvesters, the clam digging study areas were located in lower density gaper and butter clam habitat which were slightly higher in the intertidal than those commonly harvested by the public. Within this study area 40 random (1 m2 ) plots were selected for the clam digging experiment and 100 similar plots were randomly selected for the clam raking experiment. Plots were georeferenced (GPS) and marked with a numbered wooden stake located 2 m to the north of the center of each plot. 2.2. Clam raking experiment The 100 plots used in the clam raking experiment were randomly assigned as either control or treatment plots. In June, July, and August 1998, all 50 treatment plots were raked using a four tined hoe (each tine ≈20 cm in length) in a technique similar to that employed by recreational clammers to remove cockles (C. nuttalli). Immediately after the final raking
66
B.L. Boese / Aquatic Botany 73 (2002) 63–74
treatment, at least 10 eelgrass shoots were marked in 10 control and 10 treatment plots for determining leaf elongation rates using the shoot marking techniques and associated laboratory measurements as described in Kentula and McIntire (1986). Two weeks after shoot marking, these 20 plots were sampled. Plots were sampled using a 1.0 m2 quadrat. Percent plant cover was estimated by the point/intercept method. This procedure was accomplished by placing a 0.25 m2 quadrat in the south corner of the 1.0 m2 quadrat. The 0.25 m2 quadrat had an attached grid composed of 25 intersection points. Each of these intersection points was examined and the category of material (Z. marina, green algae, brown algae, red algae, or open substrate) that was observed on the surface at each intersection point was noted. The 0.25 m2 quadrat was then moved to the west plot corner and the procedure repeated clockwise around the plot until all four 0.25 m2 sections of the 1 m2 plot was completed (100 intersection points per plot). Previously marked shoots were then harvested and placed in labeled plastic bags for later leaf elongation measurements. Two cores (8 cm diameter, 5 cm deep) were then taken from each plot for macro infauna and a single core (3.4 cm diameter, 5 cm deep) for grain size and organic carbon determinations (Fig. 2A). Macro infaunal cores were transported to the laboratory and immediately sieved (0.5 mm mesh) with the retained material from the two cores combined into a single sample from each plot. At the same time the macro infauna and grain size cores were placed in the plot, four additional cores (15 cm diameter, 100 cm deep) for seagrass biomass and megafaunal measurements were also randomly placed in the plot. This coring was accomplished by placing four plastic rings (15 cm internal diameter) randomly in each plot. All seagrass leaves which were associated with shoots contained within these rings were pulled into the ring, and leaves associated with shoots which were outside of the rings were pulled out. Seagrass shoots and associated leaves within these rings were then harvested by cutting shoots at the sediment surface. Harvested material from within the four rings were combined into a single sample from each plot. A large coring device (15 cm diameter, 100 cm deep) was then positioned over each ring, the ring removed and the large core pushed into sediment to depth (approximately 1 m). Core contents were extracted and sieved (3 mm mesh). Rhizomes and associated roots retained on the sieve were separated from megafauna material and combined into a single sample for each plot for below-ground biomass determinations. Megafauna retained on the sieves were also combined from the four cores into a single composite sample from each plot. Mega- and macrofaunal samples were preserved in 10% buffered formalin then transferred into 70% ethanol for later sorting, taxonomic identification, and biomass determinations. To determine primary production (leaf elongation), new leaf material (≥2 weeks old) was separated from old leaf material (>2 weeks old) following the method of Kentula and McIntire (1986). The new growth from each plot was combined, frozen, then freeze dried for 24 h (−80 ◦ C). After freeze drying attached epiphytes were separated from leaf material by brushing with a soft bristle bush. Dry weights of removed epiphytes and new leaf growth were then determined (±0.1 g). Eelgrass shoots collected in the four cores were counted to determine shoot density then dried (24 h, 100 ◦ C) and weighed (±0.1 g) for above-ground biomass determinations. Below-ground plant samples were removed from their collection bags, placed on a 1 mm sieve then rinsed with distilled water to remove adhering sediment. Rhizome pieces with
B.L. Boese / Aquatic Botany 73 (2002) 63–74
67
68
B.L. Boese / Aquatic Botany 73 (2002) 63–74
attached roots were then removed from the sieve, dried (24 h, 100 ◦ C) and weighed (±0.1 g) for a measure of below-ground biomass. Sediment samples collected for geochemical analysis were homogenized by gentle stirring and a 5 g portion removed and analyzed for TOC using a 2400 elemental analyzer (Perkin-Elmer). The remainder of each of these samples was stored at 4 ◦ C for particle size analysis. Particle size distribution, expressed as percent sand, silt, and clay was determined using sieve-pipette methods (Buchanan, 1984). 2.3. Clam digging experiment In May 1999, 40 1 m2 plots were randomly selected in an area of the Sally’s Bend mudflat which has a resident population of gaper (T. capax) and butter clams (S. giganteus). Beginning in May, 1999, clam digging treatments were applied to 20 of these plots by pushing a 5 gal bucket (bottom removed) into the sediment at the south corner of the plot. The sediment within this coring device (≈25 cm diameter, 30 cm deep) was then removed and placed in the north corner of the plot (Fig. 2B). This digging process was repeated in June and again in July 1999 using the east and west plot corners of these same 20 treatment plots with the excavated material always deposited in the north plot corner. One month after the last of these three treatments (August 1999), 10 treatment and 10 control (undug) plots were sampled for percent cover, above- and below-ground Z. marina biomass, macrofauna, megafauna, and geochemical parameters. Sampling methodologies used for these parameters were similar to those employed for the clam raking experiments (Fig. 2A). The following year (June 2000) the remaining unsampled control and treatment plots were sampled using these same procedures with the exceptions that infaunal and geochemical samples were not taken.
3. Results 3.1. Plant measurements Each of the monthly clam raking treatments visibly removed large numbers of eelgrass leaves, some below-ground rhizomes and almost all green algae from each of the treatment plots. However, 2 weeks after the last of these monthly treatments, there were no statistically significant differences (t-tests, P > 0.05) between treatment and control plots in percent cover (Fig. 3A), Z. marina above- and below-ground biomass (Fig. 3B), Z. marina new leaf production and associated epiphytic growth (Fig. 3C). In contrast, statistically significant effects of clam digging treatments were observed (August 2000) in Z. marina above- and below-ground biomass measurements (Fig. 4) 1 month after the end of the last treatment. This trend for lower biomass in treatment plots was still apparent in June 2000, 10 months after treatments ended, however, those differences were not statistically different (Fig. 4). There were no statistically significant differences (t-tests, P > 0.05) between treatment and control plots in percent cover (Fig. 5A), Z. marina new leaf production and associated epiphytic growth (Fig. 5B). Canopy height (mean ± 2 S.E.M.) in control plots (129 ± 12 cm) was similar to that observed in treatment plots (123 ± 12 cm). Also similar in control and treatment plots were the mean number of blades
B.L. Boese / Aquatic Botany 73 (2002) 63–74
69
Fig. 3. Comparison of plant metrics in clam raking treatment and control (undisturbed) plots measured 2 weeks following final raking treatment: (A) percent cover of Z. marina and green algae by point/intercept method; (B) Z. marina above- and below-ground biomass; (C) Z. marina new leaf biomass produced in 2 weeks and associated new leaf epiphytes. Values are mean + 2 S.E.M. (error bars).
per shoot (control = 4.7 ± 0.3; treatment = 4.2 ± 0.4) and the mean width of the widest blades (control = 7.6 ± 0.6; treatment = 7.2 ± 0.8). 3.2. Benthic infauna In both clamming experiments, there was no statistical difference between control and treatment plots in the number of species or abundance of macrofauna (Fig. 6A and B).
70
B.L. Boese / Aquatic Botany 73 (2002) 63–74
Fig. 4. Comparison of above- and below-ground Z. marina biomass found in clam digging treatment and control (undisturbed) plots 1 month (August 1999) and 10 months (June 2000) after the final clam digging treatment. Values are mean + 2 S.E.M. (error bars); (∗) statistical difference from control (t-test, P < 0.05).
Fig. 5. Comparison of plant metrics in clam digging treatment and control (undisturbed) plots measure 1 month following the final clam digging treatment: (A) percent cover of Z. marina and green algae by point/intercept method; (B) Z. marina new leaf biomass produced in 2 weeks and associated new leaf epiphytes. Values are mean + 2 S.E.M. (error bars).
B.L. Boese / Aquatic Botany 73 (2002) 63–74
71
Fig. 6. Comparisons of benthic infaunal metrics between clamming treatment and control (undisturbed) plots measured 2 weeks (clam raking experiment) and 1 month (clam digging experiment) following the final clamming treatments: (A) macrofaunal species richness (number of macrofaunal species/sample) in control and treatment plots; (B) number of individual macrofaunal and megafaunal organisms/sample in treatment and control plots; (C) megafaunal biomass (g wet wt./sample) in treatment and control plots. Values are mean + 2 S.E.M. (error bars).
Megafaunal biomass and abundance was also similar between control and treatment plots (Fig. 6B and C). Dominant macrofaunal taxa in clam raking control and treatment plots were oligochaetes, two polychaetes (Mediomastus californiensis and Capitella capitata complex) and the tanaid, Leptochelia dubia. These four taxa comprised ≈90% or more of the total number of individuals per sample. There was no apparent difference in the
72
B.L. Boese / Aquatic Botany 73 (2002) 63–74
species composition between control and treatment plots. The number of species in the clam digging experimental area was greater than that in the clam raking area (Fig. 6A), with the oligochaetes, two polychaetes (M. californiensis and C. capitata), and two crustaceans (C. vulgaris and an unidentified Harpacticod) accounting for ≈90% of the individuals. As was observed in the clam raking plots, there was no apparent difference in species composition between clam digging experimental and control plots. No differences in sediment grain size or organic carbon content were found between control and treatment plots in either experiment. There were no treatment differences in the number of megafauna in either experiment. In both study areas, most of the megafaunal biomass was composed of the burrowing shrimp, Upogebia pugettenis (40–50% of total biomass) and the clam Cryptomya californiensis (30–40% of total biomass). An additional clam, Macoma nasuta, was also present in every sample. The combined biomass of these three species comprised 85–95% of the total megafaunal biomass in both these study areas.
4. Discussion Although Z. marina leaves and some rhizome material were removed by the raking treatment, 2 weeks after the final treatment no differences were detected in any measured parameter between control and treatment plots (Figs. 2 and 6). This result strongly suggests that Z. marina recovered rapidly from this type of disturbance. In addition, the intensity (repeated treatments) and spatial area (1 m2 units) of the raking disturbance used in the present study was greater than the intensity that I have observed being used by recreational clam harvesters. Thus it appears unlikely that the recreational harvest of cockles by raking significantly damages eelgrass beds. In contrast, mimicked recreational clam digging appeared to affect Z. marina above- and below-ground biomass measures determined 1 month after the final treatment. Although not statistically significant, these trends were evident 10 months after the final treatment (Fig. 4). The interpretation of this longer term effect is unclear as these latter measurements were made at the beginning of the summer growing season in Yaquina Bay and it is unknown whether the observed differences in biomass would have persisted through the growing season. Similar reductions in Z. marina biomass which occur prior to the growing season have been shown to enhance the survival of new shoots which can then contribute to the meadow’s recovery from disturbance (Olesen and Sand-Jensen, 1994). This result is similar to that of Peterson et al. (1987) who found that a commercial hand raking method and moderate clam-kicking, a commercial harvesting method in which propeller wash is used to dislodge hard clams (Mercenaria mercenaria), initially reduced seagrass (Z. marina and Halodule wrightii) biomass by approximately 25%. One year after the treatment, no differences between control and experimental areas were apparent. However, higher intensity clam-kicking reduced seagrass biomass to about half of control levels and recovery was not complete up to 4 years post-treatment (Peterson et al., 1987). As with the present study, no effect of physical disturbance was apparent on macrofaunal invertebrates, regardless of treatment intensity (Peterson et al., 1987). In contrast to these results, Brown and Wilson (1997) noted that the mimicked commercial raking for clams and bait worms on a Maine
B.L. Boese / Aquatic Botany 73 (2002) 63–74
73
mudflat significantly reduced the total number of infaunal species. However, this result is not directly comparable to the present study in that the mudflat used was unvegetated and raking treatments were done in the winter which the authors noted would reduce the possibility of recruitment and rapid recolonization (Brown and Wilson, 1997). While the digging treatment used in the present was inherently more destructive than raking, caution should be used in interpreting these results. As with the raking treatment, the clam digging used in the present study was generally more intense than I have observed being used by the general public. Recreational clam diggers usually locate an individual clam and extract it by digging a smaller hole than that dug in the present study. In addition, only about 10% of seagrass habitat in Yaquina Bay are intensively clammed by the public (Gaumer et al., 1974). Overall the results of this study indicate that recreational clamming is not a great threat to existing eelgrass beds in Yaquina Bay, however, this conclusion should be viewed cautiously. In order to reduce the possibility that study areas would be disturbed by recreational clam diggers, the study areas selected were higher in the intertidal zone than those areas which are usually clammed. Intensively clammed areas have higher clam densities and tend to have sandier and more compacted sediments than those found in my study areas. In addition, the disturbance treatments used in the present study were only applied for one summer so the effects of repeated long-term disturbance is unknown. However, eelgrass is still present in areas of Yaquina Bay which have been intensively clammed by recreational clam harvesters for decades, suggesting that long-term effects are minor.
Acknowledgements Thanks go to Faith Cole, Adar Reed, Jill Jones, Pat Clinton, Sally Noack, Mitch Vance, Alan Taylor, David Specht, Janet Lamberson, John Sewell, Judy Pelletier, John Amos, and Ignacio Eraso for assisting in this project. Thanks also to Robert Virnstein and Craig McFarlane for reviewing the manuscript. This information has been funded wholly by the US Environmental Protection Agency. It has been subjected to the Agency’s peer and administrative review, and it has been approved for publication as an EPA document. Mention of trade names or commercial products does not constitute endorsement or recommendation for use. References Brown, B., Wilson Jr., W.H., 1997. The role of commercial digging of mudflats as an agent for change on infaunal intertidal populations. J. Exp. Mar. Biol. Ecol. 218, 46–61. Buchanan, J.B., 1984. Sediment analysis. In: Holme, N.A., McIntyre, A.D. (Eds.), Methods of the Study of Marine Benthos, 2nd Edition. IBP Handbook 16. Blackwell, Palo Alto, CA, pp. 41–65. Clinton, J.D., Andorfer, J., Rose, C., Uranowski, C., Ehringer, N., 1997. Regrowth of the seagrass Thallassia testudinum into propeller scars. Aquat. Bot. 59, 139–155. den Hartog, C., 1977. Structure, function and classification in seagrass communities. In: McRoy, C.P., Helfferich, C. (Eds.), Seagrass Ecosystems: A Scientific Perspective. Marcel Dekker, New York, pp. 89–121. Everett, R.A., Rutz, R.M., Carlton, J.T., 1995. Effect of oyster mariculture on submerged aquatic vegetation: an experimental test in a Pacific northwest estuary. Mar. Ecol. Prog. Ser. 123, 205–217.
74
B.L. Boese / Aquatic Botany 73 (2002) 63–74
Fonseca, M.S., Thayer, G.W., Chester, A.J., Foltz, C., 1984. Impact of scallop harvesting on eelgrass (Zostera marina) meadows: implications for management. N. Am. J. Fish. Manage. 4, 286–293. Gaumer, T., Demory, D., Osis, L., Walters, C., 1974. 1970–1971 Yaquina Bay resource use study. Fish Community, OR, Division of Management and Research, Newport, OR, 32 pp. Griffin, K., 1997. Eelgrass ecology and commercial oyster cultivation in Tillamook Bay, Oregon. A literature review and synthesis. Tillamook Bay National Estuary Project, Garibaldi, OR, July 1997. Kentula, M.E., 1982. Population dynamics of a Zostera marina L. Bed in Netarts Bay, Oregon. Ph.D. thesis, Oregon State University, Corvallis, OR, 158 pp. Kentula, M.E., McIntire, C.D., 1986. The autecology and production dynamics of eelgrass (Zostera marina L.) in Netarts Bay, Oregon. Estuaries 9, 188–199. Levings, C.D., 1990. Strategies for restoring and developing fish habitats in the strait of Georgia–Puget sound inland sea, northeast Pacific Ocean. Mar. Pollut. Bull. 23, 417–422. Matthews, T.R., Lazar, A.C., Hunt, J.H., 1991. Aerial observations of boating impacts to shallow water grass beds in the Florida keys. Report to Florida Department of Natural Resources, Office of Marine Program and Planning, Tallahassee, FL, 11 pp. Olesen, B., Sand-Jensen, K., 1994. Biomass-density patterns in the temperate seagrass Zostera marina. Mar. Ecol. Prog. Ser. 109, 283–291. Oregon Estuaries, 1973. Division of State Lands, Salem, OR, 50 pp. Orth, R.J., 1992. A perspective on plant–animal interactions in seagrasses: physical and biological determinants influencing plant and animal abundance. In: John, D.M., Hawkins, S.J., Price, J.H. (Eds.), Plant–Animal Interactions in the Marine Benthos. Clarendon Press, Oxford, pp. 147–164. Orth, R.J., Moore, K.A., Wilcox, D.J., Fishman, J.R., 1998. Chincoteague Bay, Virginia: effectiveness of SAV sanctuary and revegetation of SAV habitat disturbed by clam dredging. Report to Virginia Marine Research Community, 19 October 1998, 6 pp. Peterson, C.H., Summerson, H.C., Fegley, S.R., 1987. Ecological consequences of mechanical harvesting of clams. Fish. Bull. 85, 281–298. Sargent, F.J., Leary, T.J., Crewz, D.W., Kruer, C.R., 1994. Scarring of Florida’s seagrasses: assessment and management. Florida Marine Research Institute Technical Report FMRI 1h/94, St. Petersburg, FL, 62 pp. Simenstad, C.A., Wissar, R.C., 1985. 13 C evidence of the origins and fates of organic carbon in estuarine and nearshore food webs. Mar. Ecol. Prog. Ser. 11, 141–152. Thayer, G.W., Wolfe, D.A., Williams, R.B., 1975. The impact of man on seagrass systems. Am. Sci. 67, 288–296. Walker, D.I., Lukatelich, R.J., Bastyna, G., McComb, A.J., 1989. Effect of boat moorings on seagrass beds near Perth, western Australia. Aquat. Bot. 36, 69–77. Zieman, J.C., 1976. The ecological effects of physical damage from motor boats on turtle grass beds in southern Florida. Aquat. Bot. 2, 127–139. Zieman, J.C., Zieman, R.T., 1989. The Ecology of the seagrass meadows of the west coast of Florida: a community profile. US Fish Wildlife Service, Biological Report 86 (7.25), Washington, DC, 155 pp.
Effects of recreational clam harvesting on eelgrass (Zostera marina) and associated infaunal invertebrates: in situ manipulative experiments Bruce L. Boese∗ US Environmental Protection Agency, Coastal Ecology Branch, Hatfield Marine Science Center, 2111 SE Marine Science Drive, Newport, OR 973656-5260, USA Received 12 March 2001; received in revised form 17 December 2001; accepted 17 December 2001
Abstract The effect of recreational clam harvesting on eelgrass (Zostera marina L.) was experimentally tested by raking or digging for clams in experimental 1 m2 plots located in a Yaquina Bay (Newport, OR) eelgrass meadow. After three monthly treatments, eelgrass measures of biomass, primary production (leaf elongation), and percent cover were compared between experimental and control (undisturbed) plots. Benthic macro (retained on 0.5 mm mesh sieve) and mega (retained on 3 mm sieve) infaunal samples were also taken to compare species number and abundances. Results indicated that clam raking did not appreciably impact any measured parameter. In contrast, clam digging reduced eelgrass cover, above-ground biomass and below-ground biomass in measurements made 1 month after the last of three monthly treatments. Although differences between control and treatment plots persisted 10 months after the last clam digging treatment, these differences were not statistically significant. Approximately 10% of the eelgrass of Yaquina Bay is subjected to recreational clamming and as this activity is generally less intense than that employed in this study, it is unlikely that recreational clamming has a major impact on eelgrass beds in the Yaquina estuary. This conclusion should be viewed with caution as multi-year disturbances were not investigated and there are differences in sediment characteristics and clam abundances between experimental sites and those sites that are intensively harvested by the public. Published by Elsevier Science B.V. Keywords: Zostera marina; Sediment; Disturbance; Clamming
1. Introduction Eelgrass (Zostera marina L.) beds, are an important habitat in Pacific northwest (PNW) estuaries. Eelgrass meadows serve as a nursery ground for juveniles of commercially ∗ Tel.: +1-541-867-5019; fax: +1-541-867-4049. E-mail address: [email protected] (B.L. Boese).
0304-3770/02/$ – see front matter. Published by Elsevier Science B.V. PII: S 0 3 0 4 - 3 7 7 0 ( 0 2 ) 0 0 0 0 4 - 9
64
B.L. Boese / Aquatic Botany 73 (2002) 63–74
important species such as herring and as a refuge for juvenile salmonids (Griffin, 1997; den Hartog, 1977; Simenstad and Wissar, 1985; Levings, 1990). Eelgrass meadows are significant sites of primary production and eelgrass shoots can be utilized directly for food by some water fowl such as the western black brant (Kentula, 1982; Griffin, 1997), and indirectly by many species via consumption of detritus (Thayer et al., 1975). Eelgrass roots stabilize the sediment (Thayer et al., 1975) and the presence of eelgrass dampens wave energy which may serve to reduce erosion and to enhance larval settlement (Orth, 1992). Because of these characteristics, species abundances in eelgrass patches are usually greater than in other estuarine habitats (Everett et al., 1995). Seagrasses may be impacted by anthropogenic physical disturbances such as dredging (Zieman and Zieman, 1989), prop scarring (Zieman, 1976; Matthews et al., 1991; Sargent et al., 1994), and boat moorings (Walker et al., 1989; Clinton et al., 1997) which directly remove biomass and alter the physical environment. Mechanical disturbances associated with commercial shellfish and bait harvesting operations have also been shown to adversely affect seagrasses (Fonseca et al., 1984; Peterson et al., 1987; Everett et al., 1995; Orth et al., 1998) and to reduce mudflat biodiversity (Brown and Wilson, 1997). Many of these studies suggested that the resulting reductions in seagrass biomass may take years to recover to predisturbance levels. One seagrass physical disturbance mechanism that has not been studied is recreational clamming. Recreational clamming is a significant activity in PNW estuaries which may have a direct impact on eelgrass and its associated biota. In Yaquina Bay, Newport, OR, gaper (Tresus capax) and butter clams (Saxidomus giganteus) are harvested by digging in lower intertidal Z. marina meadows while cockles (Clinocardium nuttalli) are often raked from the sediment surface of higher intertidal areas. Clamming activity is not limited to intertidal areas of eelgrass meadows as some clam diggers wade at low tide into 1 m of water to dig subtidal clams. Thus, this mechanical disturbance could affect eelgrass to −2 m mean lower low water (MLLW) in Oregon estuaries. As human coastal populations and associated tourism increases, the potential for physical disturbance from these recreational activities is also likely to increase. The objective of the present research was to examine the effects of one season of mimicked recreational clam digging and raking on eelgrass (Z. marina) and associated macro and mega infaunal species. 2. Methods 2.1. Site selection Yaquina Bay, the fifth largest estuary in Oregon, encompasses an area of approximately 1582 ha with 35% intertidal and 65% subtidal (Oregon Estuaries, 1973). Approximately 100 ha of this area is covered by Z. marina which is almost all intertidal (USEPA, Coastal Ecology Branch, unpublished data). Small patches of Zostera japonica are also present but are limited to a high intertidal fringe near the shore line (Fig. 1). The majority of the estuary’s eelgrass (≈85 ha) is located within the Yaquina Bay estuary’s embayment, which can be delineated into three large distinct intertidal regions: Sally’s Bend, Idaho Flat, and
B.L. Boese / Aquatic Botany 73 (2002) 63–74
65
Fig. 1. Intertidal Sally’s Bend showing extensive Z. marina meadow with some fringing Z. japonica patches located high in intertidal near shore line: (A) clam raking study site; (B) clam digging study site.
Raccoon Flat. Sally’s Bend was chosen as the site for this study, because it has the most extensive eelgrass meadows in Yaquina Bay, is relatively flat so that all experimental plots were within a narrow tide height range, and because public access is limited as the central areas of the bed are only accessible at low tide via a boat (Fig. 1). In order to further ensure that these areas would not be disturbed by recreational clam harvesters, the clam digging study areas were located in lower density gaper and butter clam habitat which were slightly higher in the intertidal than those commonly harvested by the public. Within this study area 40 random (1 m2 ) plots were selected for the clam digging experiment and 100 similar plots were randomly selected for the clam raking experiment. Plots were georeferenced (GPS) and marked with a numbered wooden stake located 2 m to the north of the center of each plot. 2.2. Clam raking experiment The 100 plots used in the clam raking experiment were randomly assigned as either control or treatment plots. In June, July, and August 1998, all 50 treatment plots were raked using a four tined hoe (each tine ≈20 cm in length) in a technique similar to that employed by recreational clammers to remove cockles (C. nuttalli). Immediately after the final raking
66
B.L. Boese / Aquatic Botany 73 (2002) 63–74
treatment, at least 10 eelgrass shoots were marked in 10 control and 10 treatment plots for determining leaf elongation rates using the shoot marking techniques and associated laboratory measurements as described in Kentula and McIntire (1986). Two weeks after shoot marking, these 20 plots were sampled. Plots were sampled using a 1.0 m2 quadrat. Percent plant cover was estimated by the point/intercept method. This procedure was accomplished by placing a 0.25 m2 quadrat in the south corner of the 1.0 m2 quadrat. The 0.25 m2 quadrat had an attached grid composed of 25 intersection points. Each of these intersection points was examined and the category of material (Z. marina, green algae, brown algae, red algae, or open substrate) that was observed on the surface at each intersection point was noted. The 0.25 m2 quadrat was then moved to the west plot corner and the procedure repeated clockwise around the plot until all four 0.25 m2 sections of the 1 m2 plot was completed (100 intersection points per plot). Previously marked shoots were then harvested and placed in labeled plastic bags for later leaf elongation measurements. Two cores (8 cm diameter, 5 cm deep) were then taken from each plot for macro infauna and a single core (3.4 cm diameter, 5 cm deep) for grain size and organic carbon determinations (Fig. 2A). Macro infaunal cores were transported to the laboratory and immediately sieved (0.5 mm mesh) with the retained material from the two cores combined into a single sample from each plot. At the same time the macro infauna and grain size cores were placed in the plot, four additional cores (15 cm diameter, 100 cm deep) for seagrass biomass and megafaunal measurements were also randomly placed in the plot. This coring was accomplished by placing four plastic rings (15 cm internal diameter) randomly in each plot. All seagrass leaves which were associated with shoots contained within these rings were pulled into the ring, and leaves associated with shoots which were outside of the rings were pulled out. Seagrass shoots and associated leaves within these rings were then harvested by cutting shoots at the sediment surface. Harvested material from within the four rings were combined into a single sample from each plot. A large coring device (15 cm diameter, 100 cm deep) was then positioned over each ring, the ring removed and the large core pushed into sediment to depth (approximately 1 m). Core contents were extracted and sieved (3 mm mesh). Rhizomes and associated roots retained on the sieve were separated from megafauna material and combined into a single sample for each plot for below-ground biomass determinations. Megafauna retained on the sieves were also combined from the four cores into a single composite sample from each plot. Mega- and macrofaunal samples were preserved in 10% buffered formalin then transferred into 70% ethanol for later sorting, taxonomic identification, and biomass determinations. To determine primary production (leaf elongation), new leaf material (≥2 weeks old) was separated from old leaf material (>2 weeks old) following the method of Kentula and McIntire (1986). The new growth from each plot was combined, frozen, then freeze dried for 24 h (−80 ◦ C). After freeze drying attached epiphytes were separated from leaf material by brushing with a soft bristle bush. Dry weights of removed epiphytes and new leaf growth were then determined (±0.1 g). Eelgrass shoots collected in the four cores were counted to determine shoot density then dried (24 h, 100 ◦ C) and weighed (±0.1 g) for above-ground biomass determinations. Below-ground plant samples were removed from their collection bags, placed on a 1 mm sieve then rinsed with distilled water to remove adhering sediment. Rhizome pieces with
B.L. Boese / Aquatic Botany 73 (2002) 63–74
67
68
B.L. Boese / Aquatic Botany 73 (2002) 63–74
attached roots were then removed from the sieve, dried (24 h, 100 ◦ C) and weighed (±0.1 g) for a measure of below-ground biomass. Sediment samples collected for geochemical analysis were homogenized by gentle stirring and a 5 g portion removed and analyzed for TOC using a 2400 elemental analyzer (Perkin-Elmer). The remainder of each of these samples was stored at 4 ◦ C for particle size analysis. Particle size distribution, expressed as percent sand, silt, and clay was determined using sieve-pipette methods (Buchanan, 1984). 2.3. Clam digging experiment In May 1999, 40 1 m2 plots were randomly selected in an area of the Sally’s Bend mudflat which has a resident population of gaper (T. capax) and butter clams (S. giganteus). Beginning in May, 1999, clam digging treatments were applied to 20 of these plots by pushing a 5 gal bucket (bottom removed) into the sediment at the south corner of the plot. The sediment within this coring device (≈25 cm diameter, 30 cm deep) was then removed and placed in the north corner of the plot (Fig. 2B). This digging process was repeated in June and again in July 1999 using the east and west plot corners of these same 20 treatment plots with the excavated material always deposited in the north plot corner. One month after the last of these three treatments (August 1999), 10 treatment and 10 control (undug) plots were sampled for percent cover, above- and below-ground Z. marina biomass, macrofauna, megafauna, and geochemical parameters. Sampling methodologies used for these parameters were similar to those employed for the clam raking experiments (Fig. 2A). The following year (June 2000) the remaining unsampled control and treatment plots were sampled using these same procedures with the exceptions that infaunal and geochemical samples were not taken.
3. Results 3.1. Plant measurements Each of the monthly clam raking treatments visibly removed large numbers of eelgrass leaves, some below-ground rhizomes and almost all green algae from each of the treatment plots. However, 2 weeks after the last of these monthly treatments, there were no statistically significant differences (t-tests, P > 0.05) between treatment and control plots in percent cover (Fig. 3A), Z. marina above- and below-ground biomass (Fig. 3B), Z. marina new leaf production and associated epiphytic growth (Fig. 3C). In contrast, statistically significant effects of clam digging treatments were observed (August 2000) in Z. marina above- and below-ground biomass measurements (Fig. 4) 1 month after the end of the last treatment. This trend for lower biomass in treatment plots was still apparent in June 2000, 10 months after treatments ended, however, those differences were not statistically different (Fig. 4). There were no statistically significant differences (t-tests, P > 0.05) between treatment and control plots in percent cover (Fig. 5A), Z. marina new leaf production and associated epiphytic growth (Fig. 5B). Canopy height (mean ± 2 S.E.M.) in control plots (129 ± 12 cm) was similar to that observed in treatment plots (123 ± 12 cm). Also similar in control and treatment plots were the mean number of blades
B.L. Boese / Aquatic Botany 73 (2002) 63–74
69
Fig. 3. Comparison of plant metrics in clam raking treatment and control (undisturbed) plots measured 2 weeks following final raking treatment: (A) percent cover of Z. marina and green algae by point/intercept method; (B) Z. marina above- and below-ground biomass; (C) Z. marina new leaf biomass produced in 2 weeks and associated new leaf epiphytes. Values are mean + 2 S.E.M. (error bars).
per shoot (control = 4.7 ± 0.3; treatment = 4.2 ± 0.4) and the mean width of the widest blades (control = 7.6 ± 0.6; treatment = 7.2 ± 0.8). 3.2. Benthic infauna In both clamming experiments, there was no statistical difference between control and treatment plots in the number of species or abundance of macrofauna (Fig. 6A and B).
70
B.L. Boese / Aquatic Botany 73 (2002) 63–74
Fig. 4. Comparison of above- and below-ground Z. marina biomass found in clam digging treatment and control (undisturbed) plots 1 month (August 1999) and 10 months (June 2000) after the final clam digging treatment. Values are mean + 2 S.E.M. (error bars); (∗) statistical difference from control (t-test, P < 0.05).
Fig. 5. Comparison of plant metrics in clam digging treatment and control (undisturbed) plots measure 1 month following the final clam digging treatment: (A) percent cover of Z. marina and green algae by point/intercept method; (B) Z. marina new leaf biomass produced in 2 weeks and associated new leaf epiphytes. Values are mean + 2 S.E.M. (error bars).
B.L. Boese / Aquatic Botany 73 (2002) 63–74
71
Fig. 6. Comparisons of benthic infaunal metrics between clamming treatment and control (undisturbed) plots measured 2 weeks (clam raking experiment) and 1 month (clam digging experiment) following the final clamming treatments: (A) macrofaunal species richness (number of macrofaunal species/sample) in control and treatment plots; (B) number of individual macrofaunal and megafaunal organisms/sample in treatment and control plots; (C) megafaunal biomass (g wet wt./sample) in treatment and control plots. Values are mean + 2 S.E.M. (error bars).
Megafaunal biomass and abundance was also similar between control and treatment plots (Fig. 6B and C). Dominant macrofaunal taxa in clam raking control and treatment plots were oligochaetes, two polychaetes (Mediomastus californiensis and Capitella capitata complex) and the tanaid, Leptochelia dubia. These four taxa comprised ≈90% or more of the total number of individuals per sample. There was no apparent difference in the
72
B.L. Boese / Aquatic Botany 73 (2002) 63–74
species composition between control and treatment plots. The number of species in the clam digging experimental area was greater than that in the clam raking area (Fig. 6A), with the oligochaetes, two polychaetes (M. californiensis and C. capitata), and two crustaceans (C. vulgaris and an unidentified Harpacticod) accounting for ≈90% of the individuals. As was observed in the clam raking plots, there was no apparent difference in species composition between clam digging experimental and control plots. No differences in sediment grain size or organic carbon content were found between control and treatment plots in either experiment. There were no treatment differences in the number of megafauna in either experiment. In both study areas, most of the megafaunal biomass was composed of the burrowing shrimp, Upogebia pugettenis (40–50% of total biomass) and the clam Cryptomya californiensis (30–40% of total biomass). An additional clam, Macoma nasuta, was also present in every sample. The combined biomass of these three species comprised 85–95% of the total megafaunal biomass in both these study areas.
4. Discussion Although Z. marina leaves and some rhizome material were removed by the raking treatment, 2 weeks after the final treatment no differences were detected in any measured parameter between control and treatment plots (Figs. 2 and 6). This result strongly suggests that Z. marina recovered rapidly from this type of disturbance. In addition, the intensity (repeated treatments) and spatial area (1 m2 units) of the raking disturbance used in the present study was greater than the intensity that I have observed being used by recreational clam harvesters. Thus it appears unlikely that the recreational harvest of cockles by raking significantly damages eelgrass beds. In contrast, mimicked recreational clam digging appeared to affect Z. marina above- and below-ground biomass measures determined 1 month after the final treatment. Although not statistically significant, these trends were evident 10 months after the final treatment (Fig. 4). The interpretation of this longer term effect is unclear as these latter measurements were made at the beginning of the summer growing season in Yaquina Bay and it is unknown whether the observed differences in biomass would have persisted through the growing season. Similar reductions in Z. marina biomass which occur prior to the growing season have been shown to enhance the survival of new shoots which can then contribute to the meadow’s recovery from disturbance (Olesen and Sand-Jensen, 1994). This result is similar to that of Peterson et al. (1987) who found that a commercial hand raking method and moderate clam-kicking, a commercial harvesting method in which propeller wash is used to dislodge hard clams (Mercenaria mercenaria), initially reduced seagrass (Z. marina and Halodule wrightii) biomass by approximately 25%. One year after the treatment, no differences between control and experimental areas were apparent. However, higher intensity clam-kicking reduced seagrass biomass to about half of control levels and recovery was not complete up to 4 years post-treatment (Peterson et al., 1987). As with the present study, no effect of physical disturbance was apparent on macrofaunal invertebrates, regardless of treatment intensity (Peterson et al., 1987). In contrast to these results, Brown and Wilson (1997) noted that the mimicked commercial raking for clams and bait worms on a Maine
B.L. Boese / Aquatic Botany 73 (2002) 63–74
73
mudflat significantly reduced the total number of infaunal species. However, this result is not directly comparable to the present study in that the mudflat used was unvegetated and raking treatments were done in the winter which the authors noted would reduce the possibility of recruitment and rapid recolonization (Brown and Wilson, 1997). While the digging treatment used in the present was inherently more destructive than raking, caution should be used in interpreting these results. As with the raking treatment, the clam digging used in the present study was generally more intense than I have observed being used by the general public. Recreational clam diggers usually locate an individual clam and extract it by digging a smaller hole than that dug in the present study. In addition, only about 10% of seagrass habitat in Yaquina Bay are intensively clammed by the public (Gaumer et al., 1974). Overall the results of this study indicate that recreational clamming is not a great threat to existing eelgrass beds in Yaquina Bay, however, this conclusion should be viewed cautiously. In order to reduce the possibility that study areas would be disturbed by recreational clam diggers, the study areas selected were higher in the intertidal zone than those areas which are usually clammed. Intensively clammed areas have higher clam densities and tend to have sandier and more compacted sediments than those found in my study areas. In addition, the disturbance treatments used in the present study were only applied for one summer so the effects of repeated long-term disturbance is unknown. However, eelgrass is still present in areas of Yaquina Bay which have been intensively clammed by recreational clam harvesters for decades, suggesting that long-term effects are minor.
Acknowledgements Thanks go to Faith Cole, Adar Reed, Jill Jones, Pat Clinton, Sally Noack, Mitch Vance, Alan Taylor, David Specht, Janet Lamberson, John Sewell, Judy Pelletier, John Amos, and Ignacio Eraso for assisting in this project. Thanks also to Robert Virnstein and Craig McFarlane for reviewing the manuscript. This information has been funded wholly by the US Environmental Protection Agency. It has been subjected to the Agency’s peer and administrative review, and it has been approved for publication as an EPA document. Mention of trade names or commercial products does not constitute endorsement or recommendation for use. References Brown, B., Wilson Jr., W.H., 1997. The role of commercial digging of mudflats as an agent for change on infaunal intertidal populations. J. Exp. Mar. Biol. Ecol. 218, 46–61. Buchanan, J.B., 1984. Sediment analysis. In: Holme, N.A., McIntyre, A.D. (Eds.), Methods of the Study of Marine Benthos, 2nd Edition. IBP Handbook 16. Blackwell, Palo Alto, CA, pp. 41–65. Clinton, J.D., Andorfer, J., Rose, C., Uranowski, C., Ehringer, N., 1997. Regrowth of the seagrass Thallassia testudinum into propeller scars. Aquat. Bot. 59, 139–155. den Hartog, C., 1977. Structure, function and classification in seagrass communities. In: McRoy, C.P., Helfferich, C. (Eds.), Seagrass Ecosystems: A Scientific Perspective. Marcel Dekker, New York, pp. 89–121. Everett, R.A., Rutz, R.M., Carlton, J.T., 1995. Effect of oyster mariculture on submerged aquatic vegetation: an experimental test in a Pacific northwest estuary. Mar. Ecol. Prog. Ser. 123, 205–217.
74
B.L. Boese / Aquatic Botany 73 (2002) 63–74
Fonseca, M.S., Thayer, G.W., Chester, A.J., Foltz, C., 1984. Impact of scallop harvesting on eelgrass (Zostera marina) meadows: implications for management. N. Am. J. Fish. Manage. 4, 286–293. Gaumer, T., Demory, D., Osis, L., Walters, C., 1974. 1970–1971 Yaquina Bay resource use study. Fish Community, OR, Division of Management and Research, Newport, OR, 32 pp. Griffin, K., 1997. Eelgrass ecology and commercial oyster cultivation in Tillamook Bay, Oregon. A literature review and synthesis. Tillamook Bay National Estuary Project, Garibaldi, OR, July 1997. Kentula, M.E., 1982. Population dynamics of a Zostera marina L. Bed in Netarts Bay, Oregon. Ph.D. thesis, Oregon State University, Corvallis, OR, 158 pp. Kentula, M.E., McIntire, C.D., 1986. The autecology and production dynamics of eelgrass (Zostera marina L.) in Netarts Bay, Oregon. Estuaries 9, 188–199. Levings, C.D., 1990. Strategies for restoring and developing fish habitats in the strait of Georgia–Puget sound inland sea, northeast Pacific Ocean. Mar. Pollut. Bull. 23, 417–422. Matthews, T.R., Lazar, A.C., Hunt, J.H., 1991. Aerial observations of boating impacts to shallow water grass beds in the Florida keys. Report to Florida Department of Natural Resources, Office of Marine Program and Planning, Tallahassee, FL, 11 pp. Olesen, B., Sand-Jensen, K., 1994. Biomass-density patterns in the temperate seagrass Zostera marina. Mar. Ecol. Prog. Ser. 109, 283–291. Oregon Estuaries, 1973. Division of State Lands, Salem, OR, 50 pp. Orth, R.J., 1992. A perspective on plant–animal interactions in seagrasses: physical and biological determinants influencing plant and animal abundance. In: John, D.M., Hawkins, S.J., Price, J.H. (Eds.), Plant–Animal Interactions in the Marine Benthos. Clarendon Press, Oxford, pp. 147–164. Orth, R.J., Moore, K.A., Wilcox, D.J., Fishman, J.R., 1998. Chincoteague Bay, Virginia: effectiveness of SAV sanctuary and revegetation of SAV habitat disturbed by clam dredging. Report to Virginia Marine Research Community, 19 October 1998, 6 pp. Peterson, C.H., Summerson, H.C., Fegley, S.R., 1987. Ecological consequences of mechanical harvesting of clams. Fish. Bull. 85, 281–298. Sargent, F.J., Leary, T.J., Crewz, D.W., Kruer, C.R., 1994. Scarring of Florida’s seagrasses: assessment and management. Florida Marine Research Institute Technical Report FMRI 1h/94, St. Petersburg, FL, 62 pp. Simenstad, C.A., Wissar, R.C., 1985. 13 C evidence of the origins and fates of organic carbon in estuarine and nearshore food webs. Mar. Ecol. Prog. Ser. 11, 141–152. Thayer, G.W., Wolfe, D.A., Williams, R.B., 1975. The impact of man on seagrass systems. Am. Sci. 67, 288–296. Walker, D.I., Lukatelich, R.J., Bastyna, G., McComb, A.J., 1989. Effect of boat moorings on seagrass beds near Perth, western Australia. Aquat. Bot. 36, 69–77. Zieman, J.C., 1976. The ecological effects of physical damage from motor boats on turtle grass beds in southern Florida. Aquat. Bot. 2, 127–139. Zieman, J.C., Zieman, R.T., 1989. The Ecology of the seagrass meadows of the west coast of Florida: a community profile. US Fish Wildlife Service, Biological Report 86 (7.25), Washington, DC, 155 pp.