Effect of various monotypic forest canopies on earthworm biomass and feral pig rooting in Hawaiian wet forests

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Forestry plantations are functioning ecosystems, and although they differ from natural ecosystems in many important ways they are governed by the same mechanisms and can provide similar ecological habitats and ecosystem services. In this sense,
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  Effect of various monotypic forest canopies on earthworm biomassand feral pig rooting in Hawaiian wet forests Noa Kekuewa Lincoln ⇑ Ngai Tahu Research Centre, University of Canterbury, Christchurch, New Zealand a r t i c l e i n f o  Article history: Received 30 May 2014Received in revised form 28 July 2014Accepted 28 July 2014 Keywords: ForestryForest ecologyFeral pigsEarthwormsHawai‘i a b s t r a c t Forestry plantations are functioning ecosystems, and although they differ from natural ecosystems inmany important ways they are governed by the same mechanisms and can provide similar ecologicalhabitats and ecosystem services. In this sense, forestry plantations can be viewed as simple forest ecosys-tems, allowing us to better isolate and understand the mechanisms that drive forest function, structure,and biodiversity. On Hawai‘i Island, 68 forest stands representing 12 species of monotypic forestry plan-tations, in addition to stands of native forest and grass pastures, were surveyed at three sites to observethe effects of monotypic canopies on earthworm biomass and occurrence of rooting by feral pigs. The can-opy species strongly influenced earthworm biomass at each site ( r  2 = 0.98, 0.99, 0.92;  p  < 0.001). Earth-worm biomass was strongly correlated to underlying soil age when examined by individual canopyspecies ( r  2 = 0.96–0.98;  p  < 0.001). Earthworm biomass was highly correlated to the occurrence of rootingby feral pigs at each site ( r  2 = 0.92, 0.94, 0.64;  p  < 0.001). Each site exhibited a different sensitivity of pigrooting in response to earthworms. Canopy and site data could thus be used to estimate total soil distur-bance by feral pigs, and inform aspects of forest management regarding soil erosion, biodiversity habitat,and hunting or trapping of feral pigs.   2014 Elsevier B.V. All rights reserved. 1. Introduction Forestry plantations are functioning ecosystems, and althoughthey differ from natural ecosystems in many important ways theyare governed by the same mechanisms and can provide similarecosystem services (Kareiva and Marvier, 2007). In this sense, for-estry plantations can be viewed as simple forest ecosystems thatallow us to isolate and better understand the mechanisms thatdrive forest function, structure and biodiversity, and thereforemay advance better management of natural, novel, and managedforests and forestry plantations for increased economic, environ-mental or social benefit. The impact of plantations on forest func-tion, such as ecosystem services and habitat development, hasbeen well studied (e.g. Ecological Society of America, 2010;Bauhus et al., 2010). Most often observations have focused on asingle forestry species relative to nearby natural or novelecosystems.The opportunity to examine how individual canopy speciescompare to each other in their influence on the subcanopy ecosys-tem is an underutilized tool in forest ecology. The species used inforestry plantations impact the understory ecosystem directlythrough physical and biogeochemical pathways, and indirectlythrough cascading ecological effects and species interactions, suchas altering predator–prey relationships, or relationships of pollina-tion or seed dispersal (e.g. Harrington and Ewel, 1997; Leiboldet al., 2004; Shea and Chesson, 2002; Strong et al., 1984). Throughthese interactive mechanisms of community ecology, canopy spe-cies may ultimately be driving seemingly unrelated aspects of for-est communities in human-established monotypic plantations.Species composition and behavior influences the usage andvalue of forests to the local community (Bauhus et al., 2010), andmay cause tensions around forest management at the local level.In Hawai‘i, feral pigs ( Sus scrofa ) define the tensions of contempo-rary forest management, which tends to be divided between thosewho value Hawaiian forests for their natural sake and those whovalue the forests as a resource to be utilized (Tummons, 2006).Feral pigs are desired by some for hunting and traditional values(Tummons, 2006), while opposed by others seeking to preservebiodiversity and watershed function (Nogueira-Filho et al., 2009).Despite the impacts, values, and tensions associated with feral pigsin Hawaiian forests, very little work has been conducted to under-stand how they interact with forest ecology to improve manage-ment, land use designation, and capture techniques (Nogueiraet al., 2007). http://dx.doi.org/10.1016/j.foreco.2014.07.0340378-1127/   2014 Elsevier B.V. All rights reserved. ⇑ Address: Private Bag 4800, Christchurch 8140, New Zealand. Tel.: +64 808 2177710. E-mail address:  nlincoln@alumni.stanford.eduForest Ecology and Management 331 (2014) 79–84 Contents lists available at ScienceDirect Forest Ecology and Management journal homepage: www.elsevier.com/locate/foreco  Following human colonization of the Hawaiian Archipelagocirca 800AD, vast amounts of environmental change occurred(e.g. Burney and Kikuchi, 2006), resulting from direct human alter-ations of the landscape but also from faunal and floral introduc-tions (Kirch, 1982). The largest early faunal introduction was thePolynesian pig, known in Hawai‘i as  pua‘a , which served as animportant source of food (Kirch, 1979; Tomich, 1986). Environ-mental changes were accelerated with the arrival of the Europeans,accompanied and succeeded by the introduction of new flora andfauna. Captain James Cook introduced European pigs to Hawai‘iduring his first voyage to the islands in 1778 (Beaglehole, 1955).Pigs have been shown to impact native flora by grazing andtrampling (Aplet et al., 1991), particularly of the large tree ferns( Cibotium  and  Sadleria  spp.) that are the dominant sub-canopy spe-cies in native mesic- to wet-ecosystems (Anderson, 1994). Pigs aredocumented in Hawai‘i as being a major disperser of several inva-sive plant species (Diong, 1982; Griffin, 1978), disturbing the soiland substrate (Aplet et al., 1991; Barret and Stone, 1983), greatlyreducing microarthropod communities (Vtorov, 1993), enhancingerosion (Browning, 2008), creating breeding habitat for introducedmosquitoes (Baker, 1975) and altering nutrient cycling in soils(Drake et al., 1989).Hawaiian pigs rely upon earthworms ( Eisenia  spp.) as an impor-tant source of protein (Baker, 1975; Griffin, 1978). At least 35 spe-cies of earthworms are described from Hawai‘i (Blakemore, 2005).Although no earthworm species are considered endemic toHawai‘i, naturalized earthworms appear in Hawai‘i as early asthe 1820‘s (Blakemore, 2002, 2005). Earthworms compose thehighest biomass among tropical soil macrofauna and play impor-tant roles in determining soil processes (Zou and Gonzalez,1997). While earthworms consist of only 1–6.2% of pigs’ stomachcontents, they are a preferred source of protein in their diet(Diong, 1982; Anderson, 1994). Based on stomach content analysis,the majority of the rooting performed by pigs in Hawaiian forests,where there exist very few tubers or other subterranean foodsources, is assumed to be in response to earthworms (Diong,1982; Anderson, 1994).During observations of forestry stands in Hawai‘i, variable lev-els of pig rooting were strikinglyapparent, leading to the investiga-tions presented in this paper. Multiple stands of various monotypicforestry test plots and control plots of native forest and pasturewere observed at three sites on Hawai‘i Island. This data was usedto examine the relationship between canopy species, earthwormbiomass, and the abundance of feral pig rooting to ask: (1) doescanopy species play a significant role in determining earthwormbiomass, (2) does canopy play a significant role in determining for-est disturbance by feral pigs, (3) how does the relationshipbetween canopy, earthworms, and feral pig disturbance changeunder the different habitats observed, and (4) how can the varia-tion in disturbance under different canopy species inform manage-ment of feral pigs in Hawai‘i? 2. Methods Observations were recorded between June and August 2009 atthree field sites on Hawai‘i Island (Fig. 1). Each field site consistedof multiple monotypic plantations established between 35 and60 years ago, Kikuyu ( Pennisetum clandestinum ) dominated pas-tures established 50+ years ago, and intact native forest dominatedby ‘O¯ hi‘a ( Metrosideros polymorpha ). While a total of 14 species areexamined, five species are well represented (three separate stands)at each site, and in addition to the broader implications this paperhas particular relevance for these species within Hawai‘i Islandhabitats: Tropical Ash ( Fraxinus uhdei ), Swamp Mahogany ( Euca-lyptus robusta ), Koa (  Acacia koa ), ‘O¯ hi‘a, and Kikuyu Grass. The sitesrepresent a range of environmental conditions, primarily the aver-age rainfall and age of underlying lava flow. All the sites are notedas supporting feral pig populations and are frequented, albeit oftenillegally, by local hunters.  2.1. Field site descriptions The Ho¯ naunau study area is located on the western slopes of Mauna Loa volcano centered at 19  29 0 21.34 00 N and155  51 0 55.82 00 W. The site occurs at an average elevation of 950 mwith an average annual rainfall of approximately 1600 mm/year,and occurs on lava flows approximately 7500 years old. The siteconsists of 20 timber stands representing 11 species, all of whichwere surveyed. Surveyed species were  A. koa, Alnus nepalensis, Cun-ninghamia lanceolata, Cupresses sempervirens, Eucalyptus globules, E.robusta, Eucalyptus saligna, F. uhdei, Pinus taeda, Sequoia sempervi-rens,  and  Toona ciliata . The site was srcinally native wet forestwith ‘O¯ hi‘a dominated canopy and Hapu‘u ( Cibotium  spp.) domi-nated subcanopy, and was cleared in the mid- to late-1960’s toestablish test plantations of commercial non-native hardwoodspecies.The Waia¯ kea site is located on the northeastern slopes of MaunaLoa volcano centered at 19  36 0 09.60 00 N and 155  08 0 00.70 00 W. Thesite occurs at an average elevation of 390 m with an averageannual rainfallin excess of 6000 mm/year,and occurs on lava flowslessthan 1500 years old. The site consists of 228 timberstands rep-resenting 12 species, of which 21 timber stands and 7 species weresurveyed. Surveyed species were  A. koa ,  A. nepalensis, E. robusta, E.saligna, Flindersia brayleyana, F. uhdei,  and  T. ciliata . The site wassrcinally native wet forest with ‘O¯ hi‘a dominated canopy andHapu‘u dominated subcanopy, and was cleared in the mid-1960sto establish test plantations of commercial non-native and nativehardwood species.The Ha¯ ma¯ kua site is located on the northeastern slopes of Mauna Kea volcano centered at 19  59 0 27.24 00 N and155  19 0 30.18 00 W. The site occurs at an average elevation of 750 mwith an annual average rainfall of approximately 2500 mm/year,and occurs on lava flows approximately 70,000 years old. The siteconsists of nine monotypic timber stands representing 3 species,all of which were surveyed. Surveyed species were  A. koa ,  E.robusta,  and  F. uhdei . The site was srcinally native wet forest withKoa and ‘O¯ hi‘a dominated canopy and Hapu‘u dominated subcano-py. All areas were logged and then cleared in the 1940–1960s toestablish pastures, with areas reforested in the 1970–1980s.  2.2. Survey methods At each site, adjacent monotypic forestry stands, undisturbednative forest and grass-dominated pasture were surveyed. Thenative forests represent baseline levels, as all sites were native for-est before being altered to monotypic forestry, while pastures rep-resents a comparative, treeless land use. Where possible, threeseparate stands of each cover type were surveyed at each site. Aline-intercept method was used to survey for pig disturbance usingtransects thatwere 100 m long and2 m wide; the transectdistancewas doubled to 200 m at the Waia¯ kea site due to low occurrence of pig disturbance. For each meter along transects, the presence orabsence of pig disturbance within the 2 m swath was noted. Fivetransects were surveyed within each stand that were situated atrandomly generated points and compass bearings, and did notapproach within 50 m of the plantation boundary. Point analysiswas used to survey for earthworms; three points were randomlylocated within each forest stand that did not approach within50 m of the plantation boundary nor another separate point. Ateach point surveyed, a one square meter soil pit was excavatedto a depth of 30 cm. Soils were sieved through one-quarter inch 80  N.K. Lincoln/Forest Ecology and Management 331 (2014) 79–84  sieve while examining for earthworms. Earthworm abundance andlength were recorded. The total length of earthworms sampled ateach point was used as a proxy for earthworm biomass. Our indi-cator, which relies only on worm length, underestimates the actualchangein wormbiomasssincelonger wormswere invariablyfatteras well. Sampling was modified at the Waia¯ kea site due to low soildevelopment; at Waia¯ kea we sampled the depression nearest tothe randomly generated point, and excavated until reaching30 cm depth or bedrock, whichever was shallower. Results wereextrapolated to 0.3 m 3 to account for differences in pit sizes.  2.3. Data analysis Transect and point-survey results for each stand were averagedand used for one-way analyses in total and by site. Effects due tocanopy species and site were examined separately using both Wil-cox and Median non-parametric test of all species, and an ANOVAusing only the five treatments common to all sites. Transect andpoint-survey results were used for linear regression examiningthe relationship between feral pig rooting and earthworm biomassin total and by site. The sensitivity of the response of feral pig root-ing to earthworm biomass by site is discussed using the slope of the linear regressions. 3. Results  3.1. Earthworm biomass A total of 68 forest-stands representing 12 canopy species wereobserved at the three sites; three forestry species were common toall sites in addition to ‘O¯ hi‘a forest and Kikuyu pastures. Earth-worm biomass differed significantly and substantially by both can-opy species and site (Fig. 2); rank ordering of species by site, or siteby species, was consistent across all cover types. When using onlythe well-represented species found at all three sites (Fig. 3), statis-tical analysis showed a significant effect of canopy (ANOVA, r  2 = 0.42,  p  < 0.001) and of site ( r  2 = 0.33,  p  < 0.001); the effect of canopy species within sites was very strong and highly significant(Ho¯ naunau  r  2 = 0.98,  p  < 0.001; Ha¯ ma¯ kua  r  2 = 0.99,  p  < 0.001;Waia¯ kea  r  2 = 0.92,  p  < 0.001). Tests were also significant when Fig. 1.  Map showing location of field sites on Hawai‘i Island, generated from Google Earth. Fig. 2.  Earthworm biomass indicator for all forest stands surveyed by site. Speciessurveyed are (A)  Sequoia sempervirens , (B)  Eucalyptus globulus , (C)  Eucalyptusrobusta , (D)  Eucalyptus saligna , (E)  Acacia koa , (F)  Meterosideros polymorpha , (G)  Alnus nepalensis , (H)  Pinus taeda , (I)  Cupressus sempervirens , (J)  Cunninghamialanceolata , (K)  Pennisetum clandestinum , (L)  Flindersia brayleyana , (M)  Fraxinus uhdei and (N)  Toona ciliata . N.K. Lincoln/Forest Ecology and Management 331 (2014) 79–84  81  using all forest stands surveyed using Wilcox and Meridiannon-parametric tests, in both grouped analysis and analysis by site(Table 1).The three sites surveyed showed vastly different quantities of earthworm biomass per volume of soil. We examined the relation-ship of earthworm biomass against average annual rainfall, eleva-tion and the age of underlying lava flow. The age of underlying lavaflow showed a highly significant relationship with earthworm bio-mass when examined by individual canopy species (Table 2), with r  2 values > 0.95 and  p  values <0.001. No significant relationshipwas observed with regard to rainfall or elevation. Although onlythree sites comprise the regression, the high significance withage of underlying flow and lack of significance with elevationand rainfall are indicative of the importance of lava flow age inthe case of this study.  3.2. Disturbance by feral pigs Based on all stands surveyed, there is a significant linear rela-tionship between pig rooting and earthworm biomass ( r  2 = 0.50,  p  < 0.001). A linear regression by site greatly strengthens the rela-tionship (Ho¯ naunau  r  2 = 0.92,  p  < 0.001, Ha¯ ma¯ kua  r  2 = 0.94,  p  < 0.001, Waia¯ kea  r  2 = 0.64,  p  < 0.001), with each site defined bya different slope (Fig. 4). When regressing only the species com-mon to each site, the overall relationship is unchanged ( r  2 = 0.50,  p  < 0.001), with the relationship being weaker for Ho¯ naunau( r  2 = 0.86,  p  < 0.001), unchanged for Ha¯ ma¯ kua ( r  2 = 0.94,  p  < 0.001), and stronger for Waia¯ kea ( r  2 = 0.78,  p  < 0.001). 4. Discussion The results from this study indicate that forest canopy specieshave a strong influence on earthworm biomass. From a commonhistory of native forest, individual monotypic canopies show a con-sistent divergence that either increases or decreases earthwormabundance in reference to native forests, and that this influenceis maintained across a large variation in rainfall and elevation. Inthis study, the age of underlying lava flow significantly affects totalearthworm biomass. However, the flows sampled in this studywere all very young (<70,000 years), and the relationship observedmay break down when including a broader range of soil ages.Increases in earthworm biomass strongly drive an increase in root-ing and soil disturbance by feral pigs within a given ecosystem. Wepostulate that this relationship is particularly strong in Hawaiianforests due to the lack of subterranean foods (i.e. there is no reasonfor pigs to root other than to harvest earthworms) and lack of ani-mal protein to scavenge (i.e. there is a high need for pigs to root forearthworms; Diong, 1982; Anderson, 1994).Changes in feral pig rooting ( D D ) in response to the changes inearthworm biomass ( D M  ) were examined for each site. The sensi-tivity of feral pig rooting in response to earthworm biomass ( D D / D M  ) changes across the sites in a way that is not clearly explainedby climatic variables. Based on personal experience and knowledgeof the areas, I postulate that broader patterns of ecological produc-tivity and controls on pig population may be driving the observedchanges in that relationship. At Waia¯ kea, where the D D / D M   (2:34)was lowest, there is a limitation of non-protein food sources forpigs that likely places a strong limit on pig populations. The naturallimitation of pig populations may make the abundance of earth-worms less significant when compared to other limitations. AtHa¯ ma¯ kua, where the  D D / D M   (2:21) was relatively low, there isan abundance of non-protein food sources for pigs but very heavypig hunting activity. The artificial suppression of pig populationsdue to hunting may make earthworms so relatively abundant(since there are fewer pigs to eat them) that the need for pigs toseek out good ‘‘worming’’ grounds is suppressed. At Ho¯ naunau, Fig. 3.  Earthworm biomass indicator for common, well represented forest standssurveyed by site.  Table 1 Non parametric tests of earthworm biomass indicator using all forest stands. Test Site  v 2 DF  p < Site Wilcox – 15.91 2 0.001Meridian – 10.7 2 0.005 Species Wilcox – 51.59 13 0.001Meridian – 47.73 13 0.001 Species by site Wilcox Ho¯ naunau 24.48 12 0.05Ha¯ ma¯ kua 13.5 4 0.01Waia¯ kea 25.33 8 0.01Meridian Ho¯ naunau 22.43 12 0.05Ha¯ ma¯ kua 11.5 4 0.05Waia¯ kea 23.42 8 0.01  Table 2 Linear regression of earthworm biomass and underlying lava flow age. Species  r  2 F p <  A. koa  0.981 367 0.001 E. robusta  0.981 363 0.001 F. uhdei  0.971 231 0.001 M. polymorpha  0.984 420 0.001 P. clandestinum  0.957 154 0.001 Fig. 4.  Relationship between earthworm biomass indicator and soil disturbancecaused by feral pigs with linear regression by site.82  N.K. Lincoln/Forest Ecology and Management 331 (2014) 79–84  where the  D D / D M   (2:5) was highest, there is an abundance non-protein food sources for pigs but very little hunting activity. Thiscombination may drive a high demand for worms and drive thehigh sensitivity observed. Ultimately, a lower pig:worm populationmay result in a weaker response of pig rooting to earthwormabundance.Despite the effect of site on the relationship between earth-worm biomass and pig disturbance, it is clear that the canopy spe-cies strongly influences earthworm biomass. Zou (1993) showsthat the earthworm biomass under a two-species canopy fallsbetween the biomass observed under each canopy separately, sug-gesting that the relative biomass of earthworms can be estimated,or at least rank ordered, if the effects of the individual canopy spe-cies are known. This indicates that results from this study can beextrapolated to better understand the distribution of earthwormdensity in forests other than monotypic plantations. Whether theextrapolations hold in forests dominated by more than two canopyspecies is unknown. Regardless, the monotypic situations can beinformative of the total earthworm biomass in Hawaiian environ-ments. For instance the native ‘O¯ hi‘a dominated native forestsshowed considerably less (26%) earthworm biomass than Kikuyudominated pastures across all sites. As native forests were clearedfor pastures, it is therefore possible that total earthworm biomassin those systems increased 3 or 4-fold. As pastures are converted tonative forest or plantations, increases or declines of total earth-worm biomass can be predicted based on the species used forreforestation.Since the canopy cover significantly alters the abundance of earthworms (which strongly correlates to feral pig rooting), itmay be inferred that the canopy cover indirectly drives the occur-rence of rooting by feral pigs. In Hawai‘i, many heavily invasivespecies have formed nearly monotypic canopies;  Eucalyptus  spp.and Tropical Ash (both involved in this study) are examples. Thesetwo highly invasive trees were observed at the extremes of theearthworm biomass continuum, with Eucalyptus species beingalmost void of earthworms and disturbance and Tropical Ash con-sistently having high earthworm biomass and pig disturbance. Theopposing effects on pig rooting that these two common invasivespecies have could inform where to apply efforts and techniquesto maximize forest management of public and private lands.Multiple impacts have been related directly to feral pig rooting.The rooting from pigs forms small pools that serve as breedinggrounds for invasive mosquitoes (Baker, 1975), which are the pri-mary vector for avian malaria and detrimental to native bird pop-ulations (Samuel et al., 2011). The rooting of pigs has also beenshown to greatly alter nutrient cycles, nitrogen in particular, whichfavors invasive plant species (Drake et al., 1989). The foragingaction of feral pigs can cause severe erosion leading to the degra-dation of watersheds, the introduction of invasive species, andthe siltation of near shore marine ecosystems (Cuddihy andStone, 1993; Browning, 2008). While rooting is only one part of soilerosion that is biologically controlled by forest canopy (e.g. treeroot systems, sub-canopy, and leaf litter all affect erosion), pigexclusion has been shown to reduce runoff levels and total sus-pended solids within runoff (Browning, 2008). Thus, using canopyspecies and climate specific relationship data to estimate levels of pig rooting can lead to better-informed management decisions.This method could be applied to planning with species-specificvegetative cover maps from remote sensing, a technique that israpidly expanding.The negative environmental impacts of increased pig rootingare also associated with an increased potential for hunting. Whilea direct correlation has not been shown, the higher occurrence of rooting likely equates to additional time spent by a pig within aparticular area. Continuing this logic, additional time spent by apig would also increase the chances of a successful hunt in thatarea, and therefore suggests that forests invaded by Tropical Ashwould be a more desirable area to allow for hunting, or a more suc-cessful area to focus on feral pig control or trapping. Conversely,opening up areas dominated by Eucalyptus species for huntingmay not result in satisfied hunters or significant pig populationcontrol efforts. 5. Conclusion Species interactions and cascading ecological effects can be bet-ter isolated and understood in forestry systems, which representsimplified forests. This study examined the relationship betweencanopy trees, earthworms, and feral pig rooting to illustrate theclear, observable patterns within these systems, and to illustratehow these patterns can influence the effectiveness of managementstrategies and social use values. Overlooking cascading relation-ships in forest management, especially ones that strongly impactenvironmental and social benefits, may lead to non-optimal man-agement decisions. This is particularly applicable in Hawai‘i, wheremost native forest canopies are dominated by a single species andinvasive trees often form novel forests with monotypic canopies.While this study’s data is derived from plantations, the findingscan be applied more broadly to forested areas dominated by afew canopy species and can therefore be informative to forestmanagers.  Acknowledgements We thank the many people who contributed to this article. Thisproject was funded by the NSF Graduate Research Fellowship Pro-gram, the Emmett Interdisciplinary Program in Environment andResources, and Stanford School of Earth Science McGee SummerResearch Grant. Adelaide Oneal, Casey Maue and Taryn Takahashicontributed to collecting and analyzing samples. References Anderson, S.P. 1994. Some environmental indicators related to feral pig activity in aHawaiian rain forest. M.Sc. thesis, Natural Resources and EnvironmentalManagement, University of Hawai‘i, Honolulu, Hawai‘i, USA.Aplet, G.H., Anderson, S.J., Stone, C.P., 1991. Association between feral pigdisturbance and the composition of some alien plant assemblages in Hawai‘iVolcanoes National Park. Vegetation 95, 55–62.Baker, J.K., 1975. The Feral Pig in Hawai‘i Volcanoes National Park. UnpublishedReport. Research Center, Hawai‘i Volcanoes National Park, Hawai‘i, USA.Barret, R.H., Stone, C.P., 1983. Managing Wild Pigs in Hawai‘i Volcanoes NationalPark. Report for Resource Management, Hawai‘i Volcanoes National Park,Hawai‘i, USA.Bauhus, J., van der Meer, P., Kanninen, M., 2010. In: Bauhus, J., van der Meer, P.,Kanninen, M. (Eds.), Ecosystem goods and services from plantation forests,Earthscan, Washington, D.C., USA.Beaglehole, J.C., 1955. In: Beaglehole, J.C. (Ed.), The Journals of Captain James Cookon his Voyages of Discovery, Hakluyt Society for the University Press,Cambridge, England.Blakemore, R.J., 2002. Cosmopolitan Earthworms – An Eco-taxonomic Guide to thePeregrine Species of the World. VermEcology, Australia, Kippax, ACT 2615.Blakemore, R.J. 2005. In: N. Kaneko, M.T. Ito (Eds.), A Series of Searchable Texts onEarthworm Biodiversity, Ecology and Systematics from Various Regions of theWorld. COE Soil Ecology Research Group, Yokohama National University, Japan.Browning, C.A. 2008. A Preliminary Examination of the Effects of feral Pigs ( Susscrofa ) on Water Quality and Soil Loss within a Hawaiian Watershed. M.Sc.Thesis. Department of Natural Resources and Environmental Management,University of Hawai‘i, Honolulu, Hawai‘i, USA.Burney, D.A., Kikuchi, W.K.P., 2006. A Millennium of Human Activity at MakauwahiCave, Maha’ulepu. Kaua’i. Human Ecol. 34 (2), 219–247.Cuddihy, L.W., Stone, C.P., 1993. Alteration of Native Hawaiian Vegetation: Effects of Humans, their Activities and Introductions. University of Hawaii Press,Honolulu, Hawai‘i, USA.Diong, C.H. 1982. Population Biology and Management of theFeral Pig ( Sus scrofa )inKipahulu Valley, Maui. PhD dissertation. Department of Zoology, University of Hawai‘i, Honolulu, Hawai‘i, USA.Drake, J.A., Mooney, H.A., di Castri, F., Groves, R.H., Kruger, F.J., Rejmanek, M.,Williamson, M. (Eds.), 1989. Biological Invasions: A global perspective. JohnWiley and Sons, Chichester, UK. N.K. Lincoln/Forest Ecology and Management 331 (2014) 79–84  83
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