Risk Analysis:  
Vessel Biofouling  
 
      
 
     
ISBN 978-0-478-37548-0 (print) 
ISBN 978-0-478- 37549-7 (online) 
  
15 February 2011 
  
 
 
  
Risk Analysis: Ve ssel Biofouling  
 
 
 
15 February 2011 
      
Approved for general release 
   
 
   
Christine Reed 
Manager, Risk Analysis 
Ministry of Agriculture and Forestry 
  
 
Requests for further copies should be directed to:  Publication Adviser MAF Information Bureau P O Box 2526 WELLINGTON  Telephone: 0800 00 83 33 Facsimile: 04-894 0300  This publication is also ava ilable on the MAF website at  
http://www.biosecurity.govt .nz/regs/im
 ports/ihs/risk  
  
© Crown Copyright - Ministry of Agriculture and Forestry 
       
i  
Contributors to this risk analysis 
 
1. Primary author/s 
 
Dr Andrew Bell Senior Adviser 
Risk Analysis, Marine MAF Biosecurity New Zealand 
Wellington 
Simon Phillips Adviser  
Risk Analysis, Marine MAF Biosecurity New Zealand Wellington 
Dr Eugene Georgiades Senior Adviser 
Risk Analysis, Marine MAF Biosecurity New Zealand Wellington 
Dr Daniel Kluza Senior Adviser 
Risk Analysis, Marine MAF Biosecurity New Zealand Wellington 
 
2. Secondary contributors 
 
Dr Christopher Denny Adviser 
Border Standards MAF Biosecurity New Zealand 
Wellington 
 
3. External peer review 
 
John Lewis 
Principal Marine Consultant ES Link Services Pty Ltd Melbourne, Victoria, Australia  Richard Piola Senior Scientist Cawthron Institute Nelson, New Zealand     The draft risk analysis has also  been internally reviewed by: 
 Liz Jones (Border Standards);  
 Justin McDonald (Post-Clearance); 
 Melanie Newfield (Risk Analysis);  
 Howard Pharo (Risk Analysis); 
 Sandy Toy (Risk Analysis).  
 The contribution of all the review ers is gratefully acknowledged.  
 
ii  
iii Contents  Page  
 
Executive summary 1  
Definitions 7 
1. Introduction 8 
1.1. Background 8 
1.2. Scope 13  
1.3. References 14  
2. Methodology 19 
2.1. Risk Management options 19 
2.2. Hazard Identification and Scoping 19 
2.3. Risk Assessment 20 
2.4. References 22  
3. Risk management options 23 
3.1. Prevention 24  
3.2. Mitigation at the border 25 
3.3. References 30  
4. Risk assessment 34 
4.1. Amphipods and Isopods  (Order: Amphipoda and Isopoda) 34 
4.2. Bacteria and Viruses 40 
4.3. Barnacles (Order: Sessilia) 43 
4.4. Bivalve Molluscs  (Class: Bivalvia) 51 
4.5. Bristleworms  (Class: Polychaeta) 61 
4.6. Bryozoans (Phylum: Bryozoa) 68 
4.7. Crabs (Infraorder: Brachyura) 75 
4.8. Echinoderms  (Phylum: Echinodermata) 83 
4.9. Flatworms  (Phylum: Platyhelminthes) 89 
4.10. Gastropods and Chitons  (Class: Gastropoda and Polyplacophora) 95 
4.11. Hydroids (Class: Hydrozoa) 103 
4.12. Macroalgae 109  
4.13. Microalgae 117  
4.14. Other Arthropods 120 
4.15. Peanut worms (Phylum: Sipuncula) 122 
4.16. Round worms (Phylum: Nematoda) 124 
4.17. Ribbon worms (Phylum: Nemertea) 126 
4.18. Sea Anemones and Corals  (Class: Anthozoa) 127 
4.19. Sea Squirts (Class: Ascidiacea) 136 
4.20. Sponges  (Phylum: Porifera) 144 
 
  
Executive summary 
The purpose of this risk analysis is to assess the biosecurity risks asso ciated with biofouling 
organisms on the hulls of vessels arriving in  New Zealand from international waters. 
 
Species within twelve of th e twenty broad taxonomic groups assessed were determined to 
present non-negligible risks to New Zealand’s core values and for which risk management 
measures can be justified. A common feature among these groups is that they are all macro-
fouling organisms with either the potential to  significantly modify ecosystem structure and 
function, or the potential to upon impact core economic valu es, such as aquaculture. 
 Based on the outcome of this risk assessment and recognising the need to  rapidly clear vessels 
at the border, it is proposed that any macro-organisms found on the hull of an arriving 
international vessel are treated as  risk organisms. This decision is based on the difficulty in 
identifying marine organisms in-situ,  and the association of macr o-fouling organisms with the 
introduction of non-indigenous species. As a resu lt, the presence of macro-fouling organisms 
can be considered to indicate the existence of a biosecurity risk which needs to be managed.   
 
Introduction 
Biofouling is a proces
s where organisms accu mulate on a clean surface that has been 
immersed in the marine environment. Biofou ling on vessel hulls thus provides a means by 
which organisms can be transferred to new loca tions. Biocidal antifouling paints, developed 
in the 1970’s, proved very effec tive against fouling organisms (in particular, those paints 
containing the active ingredient tributyltin). Therefore, the risk of introduction of non-
indigenous species by the vessel biofouling path way was thought to be mitigated. However, 
there is an increasing amount of evidence/research indicating that  this is still an important 
entry pathway. Vessel biofouling has been associated with the introduction of non-indigenous 
organisms in many parts of the world, including New Zealand.  For example, more than 65% 
of non-indigenous marine species in the waters of New Zealand, Hawai’i (USA), and Port Phillip Bay (Australia), were likely transloc ated as biofouling. In  addition, the following 
factors have provided further opportunities for the transfer of non-i ndigenous species via 
vessel biofouling:    
o Increased shipping volumes from internationa l trade mean there is an increase in the 
volume of non-indigneous species enteri ng New Zealand (i.e. increased propagule 
pressure); 
o The opening of new trade routes to New Zealand mean there is the prospect for 
biofouling organisms, not previously enc ountered,  entering and establishing in New 
Zealand; 
o The increased transit speeds of commercial vessels can lead to increased arrival of 
viable non-indigenous species as biofouling organisms are subjected to translocation 
stresses for shorter periods. 
To better understand and manage vessel biofoul ing risks in a New Zealand context, MAFBNZ 
commissioned a multi-year research survey of international vessels arriving in New Zealand.  
The objectives of the study were to determine: 
o the identity (species), or igin (native, non-indigen ous, unknown) and extent of 
biofouling occurrence on vessels;  
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  1  
2  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis o the relationship between the presence of non-indi genous species and the 
amount/extent of biofouling on a vessel; 
o the factors that influence the presence of  non-indigenous species  and the amount of 
biofouling on vessels (e.g., vessel ma intenance regime, voyage history).  
 
Findings from the MAFBNZ resear ch programme demonstrate that: 
o All major vessel types (recreational, passenge r, fishing and commercial merchant) are 
likely to have some associated biofouling; 
o Of 187 species identified in the study , over 66% were non-indigenous to New 
Zealand; of these species, 73% are not yet established in New Zealand; 
o The greater the amount of biofouling on a vessel, the higher the number of non-
indigenous species present; 
o Biofouling organisms were predominantly  arthropods (mostly barnacles), tube-
forming marine worms, bryozoans, bivalve molluscs and macroalgae; 
o Biofouling organisms generally accumulate in niche areas1, where prevention of 
biofouling is difficult, or may not be considered important; 
o Niche areas (including seachests and intern al seawater systems) pose a substantial 
biosecurity risk despite accounting for a relatively small proportion of the hull. 
Although not all non-indigenous biofouling species will necessarily pose a threat to New Zealand’s core values, it is difficult to predict the identity and impacts of future invasive species. The introduction of so me non-indigenous species to  new locations can have far-
ranging impacts on the marine environment a nd the people reliant upon it. For example, 
worldwide, introductions of non-indigenous mari ne species have impacted upon the following 
environmental, economic and socio-cultural values:  
o Direct impacts to environmental valu es by non-indigenous species include the 
recession of kelp forests. For example, in the Gulf of Maine, USA, the non-indigenous bryozoan Membranipora membranacea have formed colonies on 
indigenous kelp fronds, causing them  to become brittle and break; 
o Direct economic impacts by non-indigenous species can affect both maritime and 
non-maritime industries. The green mussel ( Perna viridis) was first identified in 
Tampa Bay, USA, in 1999. Since then dense assemblages of these bivalves now clog 
sea water intakes of power stations and desalination plants, weigh down navigational 
buoys and foul vessel hulls and engines;  
o Further, in Sweden the economic impact of the bay barnacle Balanus improvisus , 
calculated as fouling control costs, is estimated at 166-418 million Swedish krona per 
year, of which recreational vessels  accounted for approximately 74%; 
o Sea squirts such as  Ciona intestinalis , Styela clava  and Didemnum vexillum  have 
become significant fouling pests of shellfis h aquaculture.  In Nova Scotia and Prince 
Edward Island, Canada, non-indigenous sea squirts compete with stock, causing a 
reduction in size and condition and increased  mortality.  Sea squirt fouling can 
increase labour and processing costs; for example, D. vexillum  fouling of mussel 
lines in the Marlborough Sounds costs gr owers an estimated $2.2million per annum; 
                                                
 
1Niche areas are places on a vessel’s hull where protection from anti-fouling paint degr ades quickly, areas that 
are not coated in antifouling paint, and recesses that are protected from the drag created by the boat moving 
through the water.  Niche areas include propellers, ru dder shafts, bow thrusters, seachests and dry-docking 
support strips.  
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  3 o In New Zealand,  non-indigenous tubeworm species Polydora websteri  and 
P. hoplura have had negative effects on oyster aquaculture.  
o Commercial fisheries can be  indirectly affected by non-indigenous species.  The 
Asian clam Corbula amurensis2, introduced to San Francisco Bay, has altered the 
food chain to a point where commercially  harvested anchovie s no longer occur in 
local waters;  
o Non-indigenous species may also impact on economic and social values as, in the 
State of Hawai’i, the island of Maui lo ses US$20 million of tourism revenue annually 
due to the fouling of beac hes by non-indigenous algae;  
o Recreational fisheries are al so vulnerable to non-indige nous species.  For example, 
the introduction of the clam Ruditapes philippinarum2 to Venice, Italy, has been 
linked to reduced fishing catch rates, and fishers in turn  have changed their angling 
habits. 
New Zealand receives, on average, over 3,000 international vessel arrivals annually and is 
reliant on vessels to move 98% of traded goods. New Zealand has built a reputation for marine service provision, which has significan t economic benefit for the country. Vessels 
arrive throughout the year, with  a peak in activity from October to December due to the 
arrival of recreational vessels.    Different vessel types have been associated w ith different levels of risk based upon their 
physical characteristics, maintenance history, voyage characteristics and accumulation of 
biofouling.  For example, recreati onal vessels have been identified as being high risk due to 
factors such as long lay-up periods, slow and itin erant voyages, and the ma rginal benefit of in-
service maintenance to the owner.  By contrast , commercial vessels, such as container vessels, 
have rapid turn around times and significant effi ciency incentives to have clean hulls so 
potentially pose a lesser biosecurity risk, howev er, these vessels have a large surface area of 
niche areas, such as seachests, that ar e particularly vulnerable to biofouling. 
 
Current risk management 
There are as yet no m
andatory requirements in a ny country to control biofouling. As a result, 
biofouling risk management has been predomin antly coincident with  vessels using anti-
fouling systems and in-service maintenance for their own purposes. In recent years the use of 
these measures for preventing the introduc tion of non-indigenous species has been 
encouraged by MAFBNZ communication campaigns. However, as the benefits of biofouling 
management are considered to be marginal by some vessel owners, the biofouling load is 
variable between vessels. In a ddition, there are niche areas on the hull, such as seachests, 
where fouling can accumulate without noticeably impacting performance.  As a consequence, 
even apparently well-maintained vessels may ha ve biosecurity risks that require mitigation. 
 The risks associated with biofouling are beginning to be addressed at national and 
international levels. Australia is consideri ng the implementation of biofouling management 
requirements for all international vessel arriva ls, and the International Maritime Organisation 
is investigating the developm ent of guidelines containing ‘International Measures for 
Minimizing the Transfer of Inva sive Aquatic Species through Bi ofouling of Ships”. However, 
international regulation is unlikely for at least another 5 years.    
                                                
 
2Introductions of this species have occurred via ballast  water.  Although not currently known as a vessel 
biofouling species, the environmental impacts of this orga nism are representative of the potential risks posed by 
hull fouling bivalve species.  
4  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis The ongoing risks of vessel biof ouling are of immediate concer n to MAFBNZ given that the 
marine environment is a key part of many of  New Zealand’s economic, social and cultural 
values, which include aquaculture, commercial fi sheries, tourism, recreational activities and 
biodiversity: 
o New Zealand’s marine ecosystems and species  are highly diverse, and it is estimated 
that as much as 80% of the country’s nativ e biodiversity occurs in the sea.  Species 
found only in New Zealand account for 44% of  our marine biodiversity, a very high 
number that distinguishes Ne w Zealand as a global hotspot  for marine biodiversity; 
o The commercial value of New Zealand’s wild  and farmed fisheries is $1.2–1.5 billion 
annually; 
o The majority of New Zealanders live within 50 kilometres of the coastline, and many 
people swim, fish, and gather seafood from the coast; 
o Māori have a close cultural re lationship with the ocean. It is regarded as a t āonga that 
is integral to their culture and identity. Th e sea is important to tangata whenua as a 
source of food, and the mana of hapu and iwi is  still closely linked to their ability to 
provide hospitality to visitors through plenti ful kaimoana. The sea is also an important 
part of M āori spirituality and mythology. 
It is clear that New Zealand’s coastal environments3 are particularly important with respect to 
economic, social and cultural valu es, yet these areas are also the marine habitats most likely to 
be impacted upon by non-indigenous species.  Activi ties that have the po tential to introduce 
non-indigenous species are concentrated on co astal areas (e.g., shipping and boating). In 
addition, human impacts on these coastal enviro nments (e.g., pollution and sedimentation) 
can create conditions that favour the es tablishment of non-indigenous species.   
 
Methodology 
The risk analysis was undertak en following the MAFBNZ risk analysis procedures, which are 
consistent with international standards for ri sk analysis.  Risk management options, from 
published literature, are detailed with an analys is of their likely efficacy against identified 
hazards.   
 In excess of 2,000 species have been associated with the fouling of ve ssel hulls, therefore in 
order to keep the analysis manageable, twenty broad taxonomic groups were analysed.  For each group an initial hazard identification assessment determined whether a risk analysis was required.  For groups identified as potential ha zards, the following factors were assessed 
sequentially with a determination of non-negligible or negligible for each: the likelihood of entry into New Zealand, the likelihood of establishment in New Zealand, and potential 
consequences of establishment.  The anal ysis of each group concluded with the risk 
estimation and the assessment of potential management options.  
Risk assessment results 
Species within twelve of th e twenty broad taxonomic groups assessed were determined to 
present non-negligible risks to New Zealand’s core values and for which risk management 
measures may be considered. A common featur e among these groups is that they are all 
macro-fouling organisms with ei ther the potential to significantly modify ecosystem structure 
                                                
 
3The area within New Zealand’s territorial sea (extending 12 nautical mile s from the coast) and/or the extent of 
the continental shelf (a depth profile of 250 m).   
and function, or the potential to impact upon core economic values, such as aquaculture. 
While it is difficult to forecast the exact nature of impacts, it is known that introduced marine 
species can significantly alter local environmen ts leading to impacts on ecosystem services. 
Economic, environmental, social and cultural va lues at risk from these impacts include 
aquaculture, fisheries, t āonga species, iconic habitats and re creational uses of the coastal 
environment, respectively. 
 The groups identified as posing a non-negligible risk to New Zealand are: 
 Amphipods and Isopods; 
 Barnacles; 
 Bivalves; 
 Bristleworms; 
 Bryozoans; 
 Crabs; 
 Echinoderms; 
 Flatworms; 
 Gastropods; 
 Hydroids; 
 Macroalgae; 
 Sea squirts. 
 
Risk management options 
The most effective way of eliminating the en try and establishment of  non-indigenous species 
into New Zealand would be to close the border, ho wever, this is not a feasible option. Nor is it 
possible to determine an appropriate level of tolerance for vessel biofouling beyond what is 
practically achievable. Preventative measures will reduce risks prior to a vessels arrival 
however, these measures alone may not provide an adequate level of mitigation due to the 
difficulties associated with fouling prevention.  
 Given the broad range of organisms that have b een identified as biofoulers and their potential 
to be associated with each other in fouling assemblages, generic, broad-spectrum treatments 
are likely to be most effective (i.e. it is rare that biofouling assemblages are composed of an 
individual species).  Known treat ment options range from preven tative measures, such as anti-
fouling systems, to reactive containment measures, such as in-water encapsulation or land-based cleaning.  The practicalities of each opt ion vary with vessel type, size and operating 
conditions.  As the individual options assessed are not expected to fully mitigate the risk, a 
combination of treatments may be required.  The options for eradicating or managing an esta blished species in the marine environment are 
limited, difficult to perform and expensive.  For example, the eradication of the black striped 
mussel from three marinas in the Australian No rthern Territory, cost in excess of $AU2.2 
million and required the use of a biocide to k ill all life in the marinas. In New Zealand, 
incursion responses to the sea squirt S. clava and the Mediterranean fanworm ( Sabella 
spallanzanii ) have cost $2.2 million and $1 million, respectively, with a further $1.5 million 
earmarked for continuing Sabella  control.  Given the ongoing expenses associated with 
response and control efforts, preventing the establishment of non-indigenous species is 
preferable to responding to the discovery of a new incursion.  The identification of m
arine organisms in-situ  can be difficult and even expert taxonomic 
identification in the laboratory  can be challenging and time-consuming.  Therefore, based on 
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  5  
the outcome of this risk assessment and recognis ing the need to rapidl y clear vessels at the 
border, it is proposed that any macro-organism s found on the hull of an arriving international 
vessel are treated as risk organisms. This is consistent with the curr ent anti-fouling system 
technology which seeks to limit biofouli ng to no more than a slime layer. 
 
Given that fouling is most likely to accumulate  in niche areas, it is suggested that, where 
equipment is available, inspection for macro-f ouling entails a rapid vi sual inspection of the 
hull from the waterline followed by an underwat er inspection of the stern area focusing on 
any grates and appendages such as the rudder,  propeller and driveshaft.  The presence of 
macro-fouling organisms can be considered to indicate the existence of a biosecurity risk 
which needs to be managed.  Current understanding of the efficacy of inspection to characterise vessel biofouling is limited.  As a consequence, the further devel opment and validation of inspection decisions 
would enhance this understanding.  Collection of additional information on vessel char acteristics will likely aid future inspection 
efficacy. At present there are no clear indicators of biofouling ri sk, such as voyage history or 
maintenance regime, for internati onal vessels entering New Zealand. 
 To support the development of vessel risk profil es, consideration should be given to require 
vessels to supply verifiable documentation of th eir voyage history, maintenance regime and/or 
biofouling management plan. A communications pl an informing vessel owners/masters of the 
benefits of keeping accurate r ecords would support this approach. 
 
 
    
 
  
 
  
6  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
Definitions 
Anti-fouling system: a coating, paint, surface treatmen t, surface, or device that is  used on a 
vessel to control or prevent th e attachment of organisms.  
Biofouling:  the accumulation of aquatic organisms on surfaces immersed in, or exposed to, the 
aquatic environment. 
Biosecurity: the exclusion, eradication or effective ma nagement of risks posed by pests and 
diseases to the economy, e nvironment and human health. 
Consequence:  the adverse effects or harm as a result  of entry and establishment of a hazard, 
which cause environmental, economic and/or so cio-cultural values to be degraded in the 
short or long term. 
Establishment:  perpetuation into the foreseeable future  of organisms within an area of entry. 
Hazard:  any disease or organism that has the po tential to produce adverse consequences. 
Inspector:  a person who is appointed as an inspector  under section 103 of the Biosecurity Act 
(1993). 
MAF:  the Ministry of Agriculture and Forestry. 
Negligible: to be so small or insignificant as not to be worth considering. 
Niche areas: areas on a ship that are susceptible to  biofouling due to, different hydrodynamic 
forces, susceptibility to coating system wear  or damage, or being inadequately, or not, 
painted e.g. seachests, bow thrusters, propell er shafts, inlet gratings, dry dock support 
strips etc. 
Risk:  the likelihood of the occurrence and the likel y magnitude of the c onsequences of an 
adverse event. 
Transitional facility:  any place approved as a transitional fa cility in accordance with section 39 
of the Biosecurity Act (1993) for the purpose of  inspection, storage, tr eatment, quarantine, 
holding or destruction of specified types of un-cleared goods; or part of a port declared to 
be a transitional f acility in accordance with section 39. 
Vessel:  a mobile structure of any type whatsoever  operating in the ma rine environment and 
includes floating craft, fixed or floating platforms, and fl oating production storage and 
off-loading units (FPSOs). 
  
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  7  
1. Introduction 
This risk analysis assesses th e biosecurity risks associated with vessel biofouling.  These 
analyses will inform any requirements for inco ming vessels that may be developed under the 
Biosecurity Act (1993).  This analys is applies to all vessel types.  
 
1.1. BACKGROUND 
1.1.1. The biology of marine biofouling 
Any surface imm
ersed in seawater will rapidly develop a biofilm.  A conditioning film of 
organic material will adhere w ithin minutes of immersion and is rapidly colonised by single 
celled organisms such as bacteria, microalgae and cyanobacteria (Wahl 1989, Cooksey & 
Wigglesworth-Cooksey 1995, Callow & Callow 2002) .  These unicellula r recruits exude 
extrapolymeric substances (EPS) which give the biofilm its characteristic slimy texture, and thus its colloquial description of “slime layer”.  The biofilm is also referred to as primary or 
simple fouling as it is the first biological phase in th e fouling process and has little structural 
complexity.  Multicellular organisms, such as macroalgae (where light permits), barnacles, 
bryozoa and tubeworms, subsequently settle to  create a more complex fouling cover.  This 
settlement may be facilitate d by the biofilm as the bacter ia, microalgae and/or EPS are 
attractive as food, provide an at tachment substrate, or elicit  a chemotactic response from, 
recruits of macro-fouling organisms (Lam et  al. 2003, Qian et al. 2003, Qian et al. 2007).  
However, the biofilm is not essential for macro-organism recruitment.  Spores of Ulva  spp. 
for example, have been found to recruit directly to artificial surfaces th at have been “cleaned” 
prior to exposure (Joi nt et al. 2000).   
 The complexity of biofouling assemblages increa ses as macro-fouling de velops, this creates 
habitat for other organisms, including mobile sp ecies such as isopods and crabs.  If left 
undisturbed the fouling assemblage will continue  to develop toward a steady state.  On 
artificial substrates th is may not equate with the fouli ng assemblages seen in neighbouring 
natural habitats, such as rocky reefs, as the nature of the underlying substrate can influence 
recruitment and succession processes (Glasby 1999). In addition, the abundance and availability of potential recruits, inter- a nd intra-specific compet ition between settled 
organisms and the presence of predators can all in fluence assemblage structure.  For example, 
Figure 1.1 shows the fouling assemblage f ound on a marker buoy from Wellington harbour 
which is heavily fouled by the stalked ascidian Pyura pachydermatina .  In the Wellington 
region, P. pachydermatina  is common but is usually found as  isolated individuals or small 
groups at low to moderate dens ities.  Factors which may have influenced their dominance in 
this particular case include the aspect of the buoy  in relation to light, water depth and currents, 
its free floating state, the anti -fouling paint applied (and its condition), the season in which the 
buoy was deployed, the combination of materials used to construct the buoy, and the design 
of the buoy itself.  
8  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
 
 
Figure 1.1 Pyura pachydermatina  fouling on a marker buoy removed  from Wellington Harbour (Photo 
MAFBNZ). 
 
This propensity for biofouling to occur in th e marine environment has led to significant 
research effort being expended on management solutions.  The cost of managing biofouling 
varies and often no single manageme nt option is completely effec tive. In addition, there is an 
apparent “arms race” occurring between biofouling organisms and the introduction management measures, for example, some common biofouling species have developed 
biocide resistance (e.g. Hoare et al. 1995 a & b, Cassé & Swain 2006, Dafforn et al. 2008).  
Moreover, the benefits of, and therefore the im perative for, biofouling management differs. 
This, therefore, leads to different levels of mana gement on artificial stru ctures such as vessels, 
marina pontoons and wharf piles.  Mobile structur es increase the likelihoo d of biosecurity risk 
due to their ability to transfer fouling organisms beyond their natural dispersal limits.  The 
likelihood of this risk will vary depe nding upon the effort expended on biofouling 
management by the structure’s owner/operator.  
1.1.2. The impacts of translocated fouling organisms 
The translocation of fouling organism 
s has led to significant impacts in aquatic environments 
where introductions have been made.  A gl obal acknowledgement of the problem can be 
inferred from the comprehensive research revi ew presented to the International Maritime 
Organization by the correspondence group on the “D evelopment of Interna tional Measures for 
Minimizing the Transfer of Invasi ve Aquatic Species through Biof ouling of Ships” (IMO 2008).   
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  9  
10  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis Biofouling species, both native and non-indigenous4, can impact upon maritime infrastructure. 
For example, biofouling on vessels and other mari ne structures can interfere with operations, 
impose increased loading, accelerate corrosion and cause navigation buoys to sink.  For vessels, 
the most significant impact of biofouling is  on performance, as the hydrodynamic drag 
necessitates the use of more pow er and fuel to move a ship  through the water (Lewis 1998). 
Many ships and coastal industries that use pipe d seawater for engine and equipment cooling, 
fire fighting, and potable water generation ar e impacted by biofouling within the system 
components which reduce efficiency, or cause fail ures (Jenner et al. 1998).  The introduction of 
non-indigenous species can signif icantly increase these impacts and, as a result, the operating 
costs of maritime infrastructure. 
 Although biofouling has always been a problem  for mariners, new biofouling-mediated 
introductions continue to pose a threat as ship s ply interoceanic trade routes faster and more 
frequently.  Direct economic impacts by non-indigenous species can affect both maritime and 
non-maritime industries. The green mussel ( Perna viridis) was first identified in Tampa Bay, 
USA, in 1999. Since then dense assemblages of these bivalves now clog sea water intakes of 
power stations and desalination plants, weigh down navigational buoys and foul vessel hulls and 
engines (Baker et al. 2004 In Mitchem et al. 2007). In addition, the ability of some species, for 
example the tubeworms Hydroides elegans  and H. sanctaecrucis,  to now colonise the most 
common copper-based anti-fouling coatings i ndicates that biofouling management must 
continually evolve (Lewis & Smith 1991, Lewis et al. 2006). 
 Non-indigenous species can impact the envi ronment into which they are introduced by 
dominating benthic habitats, disturbing nativ e communities, and displacing local species 
(Schwindt et al. 2001, Branch et al. 2008, Hayw ard et al. 2008, Reise et al. 2008). This is 
usually due to the species life history characteristics such as  high reproductive output, mass 
settlement and rapid growth rates (Kado 2003, Lewis 2004, Casas et al. 2008). In addition, the 
introduced species are potentially freed from the lim itations of predation, pa rasites and diseases.  
 Disturbed or bare areas in the receiving environment (for example, in ports physical disturbance 
and pollution can create colonisable spaces) provid e sites for initial colonisation and subsequent 
wider dispersal (Tyrrell & Byer s 2007). In addition to physic al monopolisation of space, 
invasive biofouling species can impact nati ve species by shading, sweeping, smothering or 
consuming larvae and juveniles (Wyatt et al. 2005, Gribben & Wright 2006). Some non-
indigenous species can alter the s ubstrates or habitat architecture  thereby indirectly changing the 
physical conditions of the habitat and affecti ng other organisms (Wallentinus & Nyberg 2007, 
Sousa et al. 2009). For example, in Germany there has been a shift from native mussels to alien 
oysters (Kochmann et al. 2008), in  Australia the physical structur e of Port Philip Bay has been 
modified by the European fanworm Sabella  spallanzanii  (O’Brien et al. 2006), and in North 
America the Chinese mitten crab Eriocheir sinensis  has eroded river and lake embankments 
(Veldhuizen & Stanish 1999).  Examples of species, likely in troduced by hull biofouling, impac ting aquaculture include the 
ascidians  Didemnum vexillum (Sinner & Coutts 2003, Coutts & Sinner 2004), Ciona intestinalis 
                                                
 
4 Once translocated and established in an environment be yond their natural range, species are referred to as non-
indigenous regardless of their impact. This term is used throughout the document when not quoting a referenced 
source. However, it is noted that this terminology is not consistent in the literature (Lockwood et al. 2007).  As a 
consequence, when referring to published literature, the terminology used in that literature will be used to ensure 
we are not misrepresenting their work.   
(Robinson et al. 2006, Hawes et al. 2007) and  Styela clava  (Bourque et al. 2003), the algae  
Colpomenia peregrina (Ribera & Boudouresque 1995), the serpulid tubeworm Ficopomatus 
enigmaticus (Read & Gordon 1991), and the bivalves Mytilus galloprovincialis and Crassostrea 
gigas  (Dinamani 1971, Apte et al. 2000) .  Wild fisheries have also been impacted by the 
introduction of biofouling sp ecies including the bryozoans Membraniporopsis tubigera  and 
Biflustra grandicella (Gordon et al. 2006) and the ascidian Didemnum  sp. (Valentine et al. 
2007).  
 In general, the impacts to aquaculture includ e the loss of product, decreased product yield, 
decreased productivity and decrea sed handling efficiencies. For example, in Nova Scotia 
(Canada), the sea squirt C. intestinalis  has caused up to 50% mort ality to blue mussels ( Mytilus 
edulis ) and incurs a loss of up to C$1.86 for every 1 kg of sea squirt per m
2 (Daigle & 
Herbinger 2009). The clubbed tunicate ( S. clava ) has also been recorded to cause major 
economic impacts due to resource competition and smothering resulting in reduced mussel 
production in Prince Edward Island (Canada) (D avis & Davis 2009). In addition to decreasing 
stock marketability and yields, sea squirt foul ing can increase labour and processing costs, for 
example, for mussel aquaculture in Prince Edwa rd Island, Canada, these additional costs were 
conservatively estimated at 19% per pound to  market price (Thompson & McNair 2004).  
 
1.1.3. Vessels as a vector 
Vessels hav
e long been identified as vectors fo r the translocation of marine species (Chilton 
1910, Allen 1953, Skerman 1960, Rainer 1995).  Howe ver, the risks associated with the 
translocation of non-indigenous species in the marine environment were not widely 
recognised until the latter part of the 20th century (e.g. Carlton & Geller 1993).  Until the 
beginning of the 21st century ballast water was generally  considered the most significant 
mechanism for non-indigenous species transloc ations. This was because the advent of 
effective biocidal anti-fouling paints (in particular those tr ibutyltin-based) was thought to 
have reduced biofouling to a negligible leve l (Lewis 2004) and the risk of ballast water 
successfully translocating an en tire ecosystem was thought to be extreme (Carlton et al. 1995, 
Gollasch et al. 2002, Minton et al. 2005).  As a consequence, mitigation of ballast water 
associated biosecurity risk has received atte ntion both domestically (Im port Health Standard 
for Ships’ Ballast Water from all Countri es, MAFBNZ 2005) and internationally (IMO 
Ballast Water Management Convention 2004). 
 Since the 1990’s, a growing body of research has demonstrated that vessel biofouling has 
continued to be a significant vector for non- indigenous species translocations (James & 
Hayden 2000, Ruiz et al. 2000, Hewitt et al. 2004 , Lewis et al. 2006).  In New Zealand, 
Cranfield et al. (1998) determined that be tween 69% and 90% of known introduced marine 
organisms were probably associated with bi ofouling as a means of introduction and, more 
recently, Kospartov et al. (in press) has refined this figure as up to 87%.  Similar assessments 
in other countries have identified that 70% of marine species introduced to Hawai’i (Eldredge 
& Carlton 2002), 70% of non-nativ e coastal marine species in North America (Fofonoff et al. 
2003), 78% of non-indigenous species in Port Ph illip Bay (Hewitt et al. 2004), and 42% of 
unintentional introductions of mari ne species to Japan (Otani 2006) , are likely to have been 
introduced by hull fouling.    The type and quantity of vessel biofouling is known to vary with vessel type, maintenance 
regime, usage and location (Stubbings 1947, Skerman 1960, Rainer 1995, James & Hayden 
2000, Lewis 2004).  As a consequence, the biosec urity risk posed by hull biofouling can be 
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  11  
variable, not only between vesse ls, but also for a vessel over time and location.  However, 
there appear to be some general differences in factors between different vessel types which 
allow a division of vessel types to  be made for biosecurity purposes. For example, in a review 
of biofouling entering Australia Hi lliard et al. (2006) noted that  recreational yachts had very 
different biosecurity risk prof iles to illegal vessel s apprehended in Australian waters and, 
thus, require separate treatment.  Simila rly, James & Hayden (2000) noted considerable 
differences in the fouling of recreatio nal yachts relative to cargo vessels. 
 MAFBNZ commissioned research on the biofouling of international vessels arriving in New Zealand, which has confirmed that the incidence of biofouling is difficult to predict being a complex combination of events (Inglis et al. in prep.).  However, there do appear to be some 
factors that influence the extent of biofouling more  than others (Inglis et al. in prep.).  The key 
factors that predicte d biofouling presence in this st udy were those related to vessel 
maintenance, although other factors of simila r predictive value differed between vessel 
classes.  The other factors influencing the pr ediction of fouling on commercial and passenger 
vessels were related to design parameters, such as tonnage.  However, for recreational vessels 
factors related to voyage history, such as time in port, contributed more to the predictive 
ability of the model (Inglis et al. in prep.).  On average New Zealand receives 3,132 ± 80 (m ean ± s.d.) vessel arrivals annually (NZCS 
2008).  Arriving vessels range in size from small (averaging 15 tons) yachts to large (>100,000 tons) tankers.  Commercial vessels account  for most of the vessel traffic (Table 
1.1).    
Table 1.1.  Annual arrivals by vessel type (NZCS 2008).  Annual average (± 1 s.d.) is for the period 2003 to 
2007. 
 
Vessel Class 2003 2004 2005 2006 2007 Annual Average (± 1 s.d.) 
Container 1080 1017 1038 1204 1205 1109 ± 90 
Bulk 485 439 418 426 408 435 ± 30 
General Cargo 321 265 227 184 206 241 ± 54 
Tanker 263 274 253 240 295 265 ± 21 Commercial vessels 
Other 273 324 317 243 226 277 ± 44 
Cruise vessels 38 29 44 56 72 48 ± 17 
Fishing vessels 96 72 78 67 75 78 ± 11 
Recreational vessels 668 633 660 635 657 651 ± 16 
Slow moving and/or specialist craft 17 23 21 43 45 30 ± 13 
 
For the purposes of risk analysis MAFBNZ  has defined five classes of vessel: 
(1) recreational vessels; 
(2) commercial vessels; 
(3) cruise vessels; 
(4) fishing vessels;  
(5) slow moving and/or specialist craft (including Naval vessels). 
 
1.1.4. The New Zealand marine environment 
Terres
trial New Zealand is a collection of is lands extending over tw enty two degrees of 
latitude.  By contrast, New Zealand’s marine environment, as defined by the Exclusive 
Economic Zone (EEZ), is the fifth largest in the world stretching ove r thirty degrees of 
latitude and encompasses approximately 4.2 million km2 (about 14 times the terrestrial 
landmass) (Gordon 2009).  This area contains a diverse range of hab itats, biodiversity and 
ecosystems.   
12  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
 
New Zealand’s marine environment has been id entified as containing 80% of New Zealand’s 
biodiversity.  The north-eastern  North Island is well connected to subtropical Australian 
waters, however, the rest of New Zealand is re latively isolated, lead ing to high levels of 
endemism (Gordon 2009).  The currents around New Zealand can be coarsely described as a flow of subtropical water 
from the Tasman Sea circulating around New Zealand’s main islands before continuing 
eastwards.  However, with a similarly eastward flow of subantarctic water to the south, the 
annual average sea surface temperature ranges from 18°C in the northeast of the North Island to 11°C in the southeast of the South Island.  The combination of current and temperature 
environments has led to bi ogeographic regionalisation (Mor ton 2004) dividing the north-
eastern North Island (Auporian province) from the remaining North and entire South Islands 
(Cookian province).  The significant difference in these two provinces is derived from the 
influence of the East Australian Current, which flows down into the Auporian province, 
resulting in warmer waters and a migration of tropical species, many of which have their 
southern limit off eastern Northland (Morton 200 4).  In addition, New Zealand is exposed to 
large storm systems which, combined with the current flows, lead to well mixed coastal 
waters supporting high productivity (Laing & Chiswell 2003).    For the purpose of this risk analysis New Zeal and’s marine environment can be refined to the 
territorial sea which is  the current juri sdiction of the Biosecurity Act 1993.  Further limitation 
to the territorial sea around the dominant land masses (North Island and South Island) can be 
made as these land masses contain all of the official ports of first entry which are exposed to the greatest likelihood of entry and establishm ent and present the hubs from which any newly 
established species will disperse.  This zone is  still extensive with approximately 5,000 km of 
coastline covering environments ranging from pristine to highly modified, sheltered to 
exposed, and estuarine to oceanic.  
1.2. SCOPE 
This risk analysis considers the biosecurity ri sks from general marine biofouling on the hulls 
of vessels arriving from intern ational waters.  Biofouling ma y occur on all submerged vessel 
surfaces including internal niches, such as seachests and bow thrusters.   
 Organisms which are found exclusiv ely in ballast water, ballast  tanks, holds, anchor wells, 
bait wells, heat exchangers or any structures located above th e waterline have been excluded 
from the scope of this document. Diseases and parasites associated with biofouling organisms 
have also been excluded due to the need for fu rther scientific research confirming the extent 
and effects of the transfer of  diseases and parasites via non-indigenous marine organisms 
(Vignon & Sasal 2010), the high dive rsity of fouling organisms a nd the diseases and parasites 
that they may carry, and the time constraint s associated with this risk analysis.  
 An analysis of the options available to mana ge biofouling is included, however, no specific 
recommendations are supplied.  Different vesse l classes may have different risks thus 
analyses specific to each class will be appe nded to this analysis as they are completed. 
  
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  13  
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invasion of wave-exposed environments.  Biological Invasions  10: 103-115. 
Schaffelke, B; Hewitt, C L (2007) Impacts of introduced seaweeds.  Botanica Marina  50: 
397-417. 
Schwindt, E; Bortolus, A; Iribarn e, O O (2001) Invasion of reef- builder polychaet e: direct and 
indirect impacts on the na tive community structure. Biological Invasions 3: 137-149. 
Sinner, J; Coutts, A D M (2003) Benefit-cost analysis of management options for Didemnum 
vexillum  in Shakespeare Bay. Cawthron Report No.924. 
Skerman, T M (1960) Ship fouling in New Zealand waters: A survey of marine fouling 
organisms from vessels of the coastal and overseas trade.  New Zealand Journal of 
Science  3: 620-648. 
Smith, J E; Hunter, C L; Smith, C M (2002) Di stribution and reproductive characteristics of 
non-indigenous and invasive marine algae in the Hawai’ian Islands.  Pacific Science  
56: 299-315. 
Sousa, R; Gutierrez, J L; Aldridge, D C ( 2009) Non-indigenous invasive bivalves as 
ecosystem engineers. Biological Invasions 11: 2367-2385. 
Stubbings, H G (1947) Biology of the unde rwater fouling of ships’ hulls. Journal of Naval 
Engineering  1: 41-50. 
Thompson, R; MacNair, N (2004) An overview of the clubbed tunicate ( Styela clava ) in 
Prince Edward Island. PEI Department of Agricultur e, Fisheries, Aquaculture and 
Forestry Technical Report 234 
Tyrell, M C; Byers, J E (2007) Do artificial substrates favour  non-indigenous fouling species 
over native species? Journal of Experimental Marine Biology and Ecology 342: 54-60. 
Valentine, P C; Collie, J S; Reid, R N; Asc h, R G; Guida, V G; Blackwood, D S (2007) The 
occurrence of the colonial ascidian Didemnum  sp. on Georges Bank gravel habitat - 
Ecological observations and potential effects on groundfis h and scallop fisheries. 
Journal of Experimental Marine Biology and Ecology 342: 179-181. 
Veldhuizen, T C; Stanish, S (1999) Overview of the life history, distribution, abundance and 
impacts of the Chinese mitten crab,  Eriocheir sinensis. Department of Water 
Resources, Environmental Services Office, Sacramento, California. 
Vignon, M; Sasal, P (2010) Fish introduction and parasite s in marine ecosystems: the need for 
information. Environmental Biology of Fishes  87:1-8.  
Wahl, M (1989) Marine epibiosis. I. Fouling and anti-fouling: some basic aspects. Marine 
Ecology Progress Series  58: 175-189. 
Wallentius, I; Nyberg, C
 D (2007) Introduced marine organisms as habitat modifiers. Marine 
Pollution Bulletin  55: 323-332. 
Willan R C; Russell, B C; Murfet, N B; Moore, K L; McEnnulty, F R; Horner, S K; Hewitt, C 
L; Dally, G M; Campbell, M L; Bourke, S T (2000) Outbreak of Mytilopsis sallei 
(Recluz, 1849)(Bivalvia:Dreissenidae) in Australia.  Molluscan Research  20(2): 25-
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2. Methodology 
The m ethodology used in this risk analysis was adapted from the MAFBNZ risk analysis 
procedures (MAFBNZ 2006).  These procedures  provide a framework which is compliant 
with the requirements of the World Trade Organisation agreement on the application of 
sanitary and phytosanitary measures (WTO 1995) and the Biosecurity Act (1993).  In this 
risk analysis, the risk management options ar e presented prior to the risk assessment of 
taxonomic groupings, as hull biofouling assemb lages are typically managed as a whole 
rather than being organism/group specific.  The hazard identification and risk analysis 
sections assess the biosecurity risk with ea ch step contingent upon a determination of non-
negligible likelihood/consequen ce in the previous one. The risk analysis conducted is 
based on currently available information, howev er, it may be revised as new information 
becomes accessible. 
2.1. RISK MANAGEMENT OPTIONS 
Risk management options, in the context of th is risk analysis, are techniques used to 
control or eliminate the accumulation of biofou ling either for biosecurity or typical vessel 
management needs.  Published control met hods are reported along with an assessment of 
the efficacy and feasibility of application.  Ho wever, none of the risk management options 
are specifically recommended as per the revisi on of risks analysis procedures (10/10/2007). 
2.2. HAZARD IDENTIFICATION AND SCOPING 
Biosecurity hazards are organisms which could cause unwanted harm to New Zealand’s environmental, economic, social and cultural values.  For some commodities, extensive 
literature and formal lists of associated orga nisms are available.  However, no such list 
exists for species associated with vesse l biofouling and the information on species 
distributions present in interna tional environments, including ports and marinas, is limited.  
To construct an initial list of potential h azard organisms a review of all species known to 
be associated with hull fouling was initiated.    In reviewing known biofouling species it im mediately became apparent that a large 
number (>2,000) of species have been associ ated with vessel hulls  (Anderson et al. 2003, 
Minchin et al. 2006). As a resu lt, a grouping approach was adopted to make the analysis 
more manageable. Twenty broad taxonomi c groupings, which were loosely based on 
similarities in traits, were deemed appr opriate for the risk assessment (Table 2.1). 
 
Table 2.1.  The organism groups selected as repr esenting biofouling organisms.  Taxonomy indicates 
the resolution at which the group was examined. Taxo nomy follows Species 2000 (Bisby et al. 2009). 
 
Organism Group Kingdom Phylum Class Order 
Amphipoda Amphipods and Isopods Animalia Arthropoda Malacostraca 
Isopoda 
Bacteria    Bacteria and Viruses 
Viruses    
Barnacles Animalia Arthropoda Maxillopoda Sessilia 
Bivalves Animalia Mollusca Bivalvia  
Bristleworms Animalia Annelida Polychaeta  
Bryozoa Animalia Bryozoa   
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  19  
20  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis Organism Group Kingdom Phylum Class Order 
Crabs Animalia Arthropoda Malacostraca Decapoda 
Echinoderms Animalia Echinodermata   
Flatworms Animalia Platyhelminthes   
Polyplacophora  Gastropods and 
Chitons Animalia Mollusca 
Gastropoda  
Hydroids Animalia Cnidaria Hydrozoa  
Chromista Ochrophyta   
Chlorophyta   Macroalgae 
Plantae 
Rhodophyta   
Chromista    Microalgae 
Plantae Bacillariophyta   
Merostomata  
Malacostraca  
Ostracoda  Other Arthropods Animalia Arthropoda 
Pycnogonida  
Peanut worms Animalia Sipuncula   
Round worms Animalia Nematoda   
Ribbon worms Animalia Nemertea   
Sea anemones and corals Animalia Cnidaria Anthozoa  
Sea squirts Animalia Tunicata Ascidiacea  
Sponges Animalia Porifera   
 
Hazard identifications were initially c onducted for each of the twenty taxonomic 
groupings. This determined whether species wi thin the groupings represented a potential 
hazard to New Zealand via the vessel bi ofouling pathway. The hazard assessment 
addressed the following areas of interest in paragraph order: 
 broad group taxonomy, number of species and general habitats; 
 life-cycle(s); 
 ecology and importance;  
 known introduced species and their impacts. 
From the above assessment, species within th e taxonomic group were determined to be a 
potential hazard if they met the following criteria: 
 are known components of mari ne biofouling assemblages5; 
 are known to have been introduced to new locations; 
 have known impacts on core values. 
If these criteria were met, then a risk assessment was performed.  
2.3. RISK ASSESSMENT 
Risk assessment is the evaluation of the li kelihood and consequences to environmental, 
economic, social and cultural values. The ai m of the risk assessment was to identify 
hazards which presented an unacceptable level of risk, for which risk management options would be required. The overall risk estimation was based on three inte r-related evaluations: 
 the entry assessment; 
 the establishment assessment;  
 the consequence assessment. 
 
                                                
 
2 Note: The criteria “are known components of marine biofouling assemblages” does not  specifically  indicate 
that members within the group are known to foul vesse l hulls. This criteria was designed to establish whether 
or not species within the group are part of biofoulin g communities in the general marine environment. From 
this pool of general marine fouling organisms, a smaller group may be capable of fouling vessel hulls. The 
presence/absence of this smaller group was, wh ere necessary, covered by the entry assessment. 
  
As previously stated, border interception data  and in place biosecurity measures support 
the risk assessments of many commodity pathwa ys. As a result, risk descriptors may be 
used to indicate a relative leve l of likelihood, consequence, or risk. However, in the case of 
vessel hull biofouling, no border interception data and no specific measures exist, 
therefore, general information regarding organism life-historie s and literature specific to 
the subject of biofouling were used for the ri sk assessment. Uncertainties and assumptions 
encountered throughout the risk assessment were identified within the appropriate section.  
 
The entry assessment addressed the following areas of concern: 
 occurrence on pathways of vessels entering New Zealand; 
 association with vessel biofouling; 
o associations with other taxa; 
o antifouling resistance; 
 oceanic survival/tolerance; 
 recent records of entry into New Zealand. 
From this information the determination of  likelihood of entry was made based on the 
following criteria: 
 are known to occur along vessel pathways; 
 are known to be associated with vessel biofouling; 
 are known to survive translocation. 
If these criteria were met, it was determined that there was a non-negligible risk of species within the taxonomic grouping en tering New Zealand via hull biofouling. As a result, an 
establishment assessment was performed.   The establishment assessment addressed the following areas of concern: 
 suitability of the New Zealand environment; 
 reproductive characteristics; 
o dispersal characteristics; 
o propagule pressure; 
 records of recent establishment. 
From this information the determination of likelihood of establishment was made based on 
the following criteria: 
 are known to inhabit similar climates/env ironments to those that exist in New 
Zealand; 
 are capable of forming establishe d populations in New Zealand; 
 have been recently recorded as  established in New Zealand. 
If these criteria were met, it was determined that there was a non-negligible risk of species within the taxonomic groupi ng establishing in New Zealand following entry via hull 
biofouling. As a result, a conseque nce assessment was performed.  
 The consequence assessment summarised known existing and potential impacts to the four 
core values designated by MAFBNZ New Z ealand. As New Zealand data regarding the 
impacts of non-indigenous species are relativ ely scarce, internati onal impact data was 
sourced and related back to New Zealand c onditions where relevant. The examples of 
impacts within the consequence assessments , although not necessarily attributed to 
organisms introduced to those areas via vesse l hull fouling, are repr esentative of the 
potential consequences posed by species within the broad taxonom
ic groups that may be 
translocated via vessel hulls.  
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  21  
22  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis The risk estimation summarised the entry, es tablishment and consequence assessments to 
conclude whether or not orga nisms within the taxonomic gr oup presented a non-negligible 
risk to New Zealand’s core values. Possible risk mitigation measures  were outlined when 
the outcome of the risk estimation was non-negligible. 
 A precautionary approach to the application of risk management measures was adopted 
when there was significant uncertainty in th e estimated risk. In these circumstances, the 
measures proposed were consistent with thos e groups where equivalent  uncertainties exist 
and should be subject to review as soon as  additional information becomes available. 
 
2.4. REFERENCES 
Anderson, C; Atlar, M; Callow, M; Candrie s, M; Milne, A; Townsin, R L (2003) The 
development of foul-release co atings for seagoing vessels. Proceedings of the 
Institute of Marine Engineering, Scie nce and Technology Part B: Journal of 
Marine Design and Operations 4: 11-23. 
Awad, A; Clarke, C; Greyling, L; Hilliard, R; Polglaze, J; Raaymakers, S (2003) Ballast 
water risk assessment, Port of Saldanha Bay, Republic of South Africa, November 
2003: Final report . Globallast Monograph Series No. 13; IMO, London. 
Bisby, F A; Roskov, Y R; Orrell , T M; Nicolson, D; Paglinawan, L E; Bailly, N; Kirk, P 
M; Bourgoin, T; Baillargeon, G (Eds) (2009). Species 2000 & ITIS Catalogue of 
Life: 2009 Annual Checklist . Digital resource at www. catalogueoflife.org/annual-
checklist/2009/. Species 2000: Reading, UK. 
Coutts, A D M (2002) A biosecurity investigation of a barge in the Marlborough Sounds .  
Cawthron Report 744. July 2002. 
Inglis, G J; Floerl, O; Unwi n, M; Ponder-Sutton, A; Seawar d, K; Kospartov, M; Bell, A; 
Kluza, D (in prep.). The biosecurity risks associated with biofouling on international 
vessels arriving in New Zealand. Summary of the patterns and predictors of fouling.  
Biosecurity New Zealand Technical Paper. 
Kelleher, G; Bleakley, C; Wells, S (Eds) (1995) A global representative system of marine 
protected areas .  Vol I – IV.  The World Bank; Washington DC, USA. 
MAFBNZ (2006) Risk analysis procedures. Version 1. MAF Biosecurity New Zealand, 
Wellington.  
Minchin, D; Floerl, O; Savini, D; Occhipinti-Ambrogi, A (2006) Chapter 6: Small craft 
and the spread of exotic species . In Davenport, J; Davenport, J L (Eds) The ecology 
of transportation: Managing mobility for the environment. Springer; The Netherlands. 
NZCS (2008) International vessel arrival and departure information 2000 – 2007. Supplied 
by the intelligence section of New Ze aland Customs Service in March 2008. 
WTO (1995) Agreement on the application of sanitary and phytosanitary measures . World 
Trade Organisation; Geneva. 
   
3. Risk management options 
Vessel biofouling poses a biosecurity risk through the introduc tion of non-indigenous 
species which may have negative impacts on New Zealand’s environmental, social, 
cultural or economic values.  The most effectiv e way of eliminating such risk would be to 
close the border.  However, this is not a feasible option because: 
 New Zealand is reliant on shipping fo r 98% of international trade by volume 
(Statistics New Zealand 2006); 
 there is a significant additional economic be nefit associated with vessel arrivals, 
including: 
o tourism; 
o refitting; 
o resupply; 
o port services;  
 New Zealand has international obligations to  provide a port of re fuge to vessels in 
distress or attempting to escape conditi ons which may endanger life at sea; 
 there is likely to be significant discon tent from both domestic and international 
stakeholders if freedom of movement is curtailed for little obvious gain. 
Given the continual arrival of vessels, it is inevitable that non-indigenous species 
associated with biofouling on vessel hulls will enter New Zealand.  
 Due to a lack of specific understanding of th e biology of most biofouling species and the 
level of consequence they pose,  it is not currently possible to determine an appropriate 
level of tolerance for vessel biofouling beyond wh at is practically achievable. As a result, 
generic options available to minimise or elim inate biofouling to an acceptable level have 
been identified.  No analysis of the risk ma nagement objective or treatment preference(s) 
of any of the management measures is provided.    Risk management can be subdivided into preventative measures and border treatments. Preventative measures will reduce likelihood of translocation prior to  a vessels’ arrival, 
however, this option alone may not provide an ad equate level of risk mitigation due to the 
difficulties of fouling prevention.  The practical point of intervention for vessel biofouling is where the fouling exceeds a slime layer and becomes complex.  The rationale for this is: 
 the creation of a slime laye r (bacteria, viruses, microalgae) occurs rapidly on 
immersion and, aside from continuous clean ing, there is no effective preventative 
technology; 
 effective technology to prevent macro-f ouling is available and commonly used; 
 the greater the density and diversity of fouling, the more lik ely it is that non-
indigenous species will be translocated; 
 the biogeography of micro-fouling sp ecies is still widely debated. 
 The potential treatment options are described along with an analys is of their possible 
efficacy against identified hazards and any resi dual biosecurity risks. The use of measures 
to mitigate the risk from specific organism gr oupings is discussed, where applicable, in the 
specific risk management sections of those groupings. 
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  23  
3.1. PREVENTION 
3.1.1. Vessel maintenance 
The im
pact of biofouling on the operationa l performance of vessels is well known 
(Redfield & Ketchum 1952, Callow & Callow 2002).  As a consequence, most vessel 
owners make active attempts to minimise hull biofouling, with the most common management measure being the use of anti-f ouling paint.  Anti-fouling paints can be 
coarsely classified into hard, ablative, self -polishing and foul-releas e categories (Inglis et 
al. 2008).  All but the foul-release paints use bi ocides, such as copper, to deter fouling, 
however, the effectiveness of these systems can vary due to: 
 the stochastic nature of biofouling (Wahl 1981);  
 the presence or absence of species tolerant to the an ti-fouling systems (e.g. copper 
tolerant bivalves (Hoare et  al. 1995 a, b) and algae (R ussell & Morris 1970, Reed & 
Moffat 1983));  
 the type, age and condition of the anti-f ouling coating (Piola & Johnston 2008);  
 voyage and vessel characte ristics (Lewis 2004).   
 The use of in-water cleaning to maintain performance, enhance aesthetics and attempt to extend the life of the anti-f ouling system has been found to be common for recreational 
vessels but uncommon for commercial vessels (Inglis et al. in pr ep.).  By contrast, 
Bohlander (2009) noted that it is common for commercial vesse ls to complete in-service 
maintenance procedures such as propeller pol ishing and, where biofou ling is thought to be 
problematic, in-water cleaning.  In many jurisd ictions it is difficult to undertake in-water 
cleaning due to concerns rega rding non-indigenous species in troductions and the entry of 
excessive concentrations of anti- fouling biocides into the environment.  However, there is 
a developing understanding that in-service maintenance practices, such as in-water 
cleaning, could have more beneficial outcomes than risks.  For example, a review of the ANZECC code of practice for in-water hull cl eaning is currently be ing undertaken to 
examine under what circumstances the benefits may exceed the risks.   For most vessel owners the efficacy of an  anti-fouling system and any in-service 
maintenance program is measured in terms of vessel performance.  While this almost certainly reduces the potential biofouling load, fouling may accumulate where it is 
considered to not significantly impact on perf ormance and/or cannot eas ily be treated, such 
as the interiors of seachests.  From a biosecur ity perspective, the pres ence of biofouling is 
of concern regardless of the location.    The MAFBNZ commissioned research on vessel biofouling shows that fouled recreational 
vessels were most likely to have substantial macro-fouling in niche areas, such as the 
propeller and shaft, while commercial vesse ls accumulated most fouling around seachest 
grates and bow thrusters (Inglis et al. in prep.).  These finding s are similar to those of the 
Australian Shipowners Association (ASA 2006)  which noted that th e anti-fouling systems 
in niche areas are likely to be compromised or absent. In addition, water flows within niche areas may not be laminar and thus create a sa nctuary from biocides and physical forces 
which may otherwise control foul ing (ASA 2006).  Therefore, even if attention is paid to 
niche areas during in-water cleaning, they may be  more susceptible to reinfestation as the 
deterrents to settling organisms may be  either compromised or absent.   
24  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
3.2. MITIGATION AT THE BORDER 
While it is likely that a combination of an  effective anti-fouling system and good vessel 
husbandry will reduce biofouling risk, there is like ly to be a latent risk associated with 
niche areas.  Given that some vessels will not be as well maintained as they could be, it is 
necessary to consider a range of options to mitigate biofouling when the vessel is inspected 
at the border.  At present, all vessels are inspected predominantly for hazard organisms or risk goods 
carried on or in the vessel, however, hull biofouling is not generally considered as a 
primary option. Inspectors do have the ability to demand that a vesse l be hauled out and 
cleaned if they suspect that the biofouling is significan t enough to constitute a likely risk to 
New Zealand. However, the only method availabl e for the inspector to judge the level of 
fouling is observation of the hull from the dock or a pontoon. This method has obvious difficulties for the determination of fouling within vessel niche areas. Trials on the use of 
underwater pole-cameras to examine the bi ofouling on recreationa l vessel hulls are 
currently underway. 
3.2.1. Inspection 
Identifying m
arine organisms in-situ can be difficult and even expert taxonomic 
identification in the laborato ry can be challenging and time-consuming. In addition, the 
introduction of non-indigenous species has been  associated with the presence of macro-
fouling organisms (Inglis et al. in prep.). Ba sed on the outcome of this risk assessment and 
recognising the need to rapidly clear vessels at the border, it is proposed that any macro-
organisms found on the hull of an arriving in ternational vessel ar e treated as risk 
organisms. This is consistent with the curr ent preventative anti-f ouling system technology 
which seeks to limit biofouling to no more than a slime layer. 
 It is often difficult to detect the presence of fouling organisms from the surface due to 
factors such as water clarity, wharf access and accessibility of niche areas. While the 
presence of waterline fouling indicates more substantial fouling below the waterline, the 
absence of waterline fouling does not indicate the absence of substantial fouling elsewhere 
on the hull (IMO 2007). Given that fouling is most  likely to accumulate in niche areas, it is 
suggested that, where equipment is available,  inspection for macro-fouling entails a rapid 
visual inspection of the hull from the waterline followed by an underwater inspection of the stern area focusing on any grates and a ppendages such as the rudder, propeller and 
driveshaft.  The presence of macro-fouling or ganisms can be considered to indicate the 
existence of a biosecurity risk which needs to be managed.  Current understanding of the efficacy of inspection to characterise vessel biofouling is limited.  As a consequence, validation of in spection decisions needs further development. 
This can be achieved by auditing the initial ve ssel inspection decision at the border against 
a thorough inspection of biofouling (Hilliard et al. 2006). The auditing process should 
include a random inspection of vessels deemed to pose a negligible risk as well as all 
vessels directed to treat their bi ofouling (Hilliard et al. 2006).  
 The colle ction of additional information on vessel characteristics will likely aid future 
inspection efficacy. At present there are no clear indicators of biofouling risk, such as 
voyage history or maintenance regime, for international vessels entering New Zealand. 
One reason for this may have been the re latively small sample size of the MAFBNZ 
commissioned research on vesse l biofouling which targeted 15% of incoming vessels. It 
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  25  
may be that with further information better pr edictors of risk will become evident. To 
support the development of vesse l risk profiles, considerat ion should be given to the 
requirement of vessels to supply verifiab le documentation of their voyage history, 
maintenance regime and/or biofouling ma nagement plan. A communications plan 
informing vessel owners/masters of the benefits  of keeping accurate records may be put in 
place to support this approach. 
 
3.2.2. In-water treatments 
In-wate
r treatments refer to those treatments  carried out while the vessel remains in the 
water.  This may occur at the quarantine wharf or require moving the vessel to a specific 
location designated for such activity. 
3.2.2.1.  Removal by hand 
In-water removal of biofou ling organism
s by hand is common on smaller vessels and is 
generally carried out by a diver using tools, such as paint sc rapers (Floerl et al. 2005c).  
This is an effective method for removing some  organisms, especially where they are in 
isolated patches and can be contained with a low risk of viable pr opagules being released. 
For example, this approach was successfully used in the Lyttelton port in response to the 
entry of Sabella spallanzanii .   
 
While removal by hand may be effective in  removing visible organisms, it may not 
eliminate the risk.  For example, Floerl et al. (2005c) found that 72% of organisms 
removed from vessel hulls by hand scraping re mained viable after removal. Similarly, 
Woods et al. (2007) recorded 72% (summer) and 66% (winter) of removed organisms (excluding tubiculous polychaetes) as still viable one hour after in-water hand scraping.  
As a consequence, to be effective as a bi osecurity treatment opti on, hand removal must 
contain all removed organisms a nd any potential propagules.  This is most likely to be 
possible where the organisms in question ar e large and easily removed. However, the 
success of this procedure may be reduced as fouling assemblages: 
 become more complex; 
 contain organisms likely to re lease propagules during removal; 
 contain a larger proportion of organi sms strongly affixed to the surface; 
 are present in difficult to access areas.   
3.2.2.2.  Mechanical cleaning 
Mechanical in-water cleaning techn
ologies rang e from water blasters to automated rotating 
brush systems.  Generally, these systems are used on larger commercial vessels due to 
factors including:  
 expense (both to purchase and operate); 
 availability (often only a si ngle operator in a region); 
 potential to damage the hull (in particular composite hulls where fibre damage or 
delamination may be induced); 
 ineffectiveness over curved surfaces (most mechanical systems are designed to cover the hull of larg e vessels where most panels are flat or gently curved).   
 Rotating brushes are the most common mechan ical cleaning systems and have typically 
been developed for naval or commercial vessels .  However, as yet, none of the in-water 
systems currently available cap ture 100% of debris remove d during the cleaning process 
(Bohlander 2009). Moreover, they are not usually effective at cleaning the entire hull due 
26  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  27 to an inability to conform to curves and/or access areas with appendages, such as around 
the vessels stern (Floerl et al. 2005c, Hopkins et al. in press).  In-water water blasting could 
avoid some of these issues, however, the equi pment is specialised, difficult to obtain and 
difficult to use safely.  It has also been found that a significant proportion of organisms 
dislodged from the hull by in-water wate r blasting will survive (MAFBNZ 2009). 
3.2.2.3.  Encapsulation 
This m
ethod uses an impervious material to cover a vessel hull in or der to reduce the water 
volume surrounding the hull thereby creating toxi c conditions for the fouling organisms.  
Several studies have found this to be an effective method of treating fouling organisms 
(McEnnulty et al. 2001, Coutts & Forrest 2007) .  There are two methods commonly used 
for encapsulation. The first is where an appropr iate material (commonly silage wrap) is 
wrapped around the vessel to provide a tempor ary, single use, cover. The second method is 
where a shroud, suspended beneath inflatable  pontoons, is drawn around the vessel. Such 
systems are usually custom built and reusable (e.g. http://www.seapen.com.au  ).  In both 
cases the water between the encap sulation and the hull is evacua ted to expedite the process 
of attaining anoxia. It is possible to furt her accelerate the pro cess by the addition of 
freshwater or chemicals, such as acetic acid or hypochlorite. 
 The chief benefits of encapsulation are the i ndiscriminate killing of fouling organisms such 
as those in hard to access niche areas, including internal seawater systems, and the ability to rapidly deploy in most locations.  The de ployment of encapsulation technology can be 
especially rapid where reusable systems are available (i.e. manoeuvre the vessel into the 
shroud).  The availability of equipment and personnel to deploy an encapsulation method will determine the feasibility of use in most locations.  The effective use of encapsulation as a trea tment option requires further development in 
order to be able to specify the following:  
 effective kill times; 
 effective dosages for additives; 
 the appropriate use of any additives (environmental and anti-fouling system 
concerns); 
 safe containment and disposal of debris; 
 the latent risk of vessel fouling
6 (e.g. barnacle tests).   
3.2.2.4.  Heat treatment 
Heat has been used as a biofouling contro l m
easure for many different applications 
including the internal piping systems of power plants (Raj agopal et al. 1994).  To be 
effective, heat treatments need to achie ve temperatures exceeding 45°C throughout the 
entire fouling assemblage for an appropriate period of time7 (Wallentinus 1999, 
McEnnulty et al. 2001). However, for effectiv e killing the heat must be evenly applied 
                                                 
6 While fouling organisms are killed during treatment their calcareous tests can remain attached to the vessel 
hull. Once the encapsulation is removed and the hull is exposed to open sea-water again the tests provide an 
attachment site for fouling organisms and facilitate the growth of biofouling. The latent risk comes from 
larvae, spores, and organisms present in the area settling upon or sheltering in the tests and being 
subsequently translocated. Th e presence of vacated tests or dead fouling structures can be an important vector 
for the secondary spread of established non-indigenous species. 
7 As the temperature increases the exposure time needed  to kill respective organisms decreases. For example, 
to make cysts of the dinoflagellate Gymnodinium catenatum  inactive they can be exposed to water of 38-
40°C for 120 seconds or water of 45-47°C for 30 seconds (Bolch & Halegraeff 1993 In  McEnnulty et al. 
2001).  
28  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis which is difficult for dense assemblages.  Difficulties may also arise due to the large 
thermal buffering capacity of seawater, the natu re of the substrate and the nature of the 
fouling itself.  Whilst engineering solutions can overcome some of these difficulties, uniform application over dense assemblages is difficult to manage.    On vessel hulls, the only reported use of h eat treatment was for the removal of an Undaria 
pinnatifida  infestation on a sunken vessel (Wooten et al. 2004).  While successful, this 
method was found to be resource intensive a nd expensive.  A developing technology is the 
use of heat treatment for in-service maintenance.  That is, heat treatment applications are made to eliminate newly settled fouling organisms to prevent the establishment of complex fouling assemblages ( http://www.commercialdiving.com.au
).   
 
3.2.2.5.  Freshwater 
Imm
ersion in freshwater has been suggested as a biofouling treatment  by, for example, the 
sailing of a vessel into a freshwater river environment.  However, due to the risks of 
organisms releasing or spawning while being moved from marine to fresh waters, this method appears to be more appropriate for pr e-departure, i.e. a vessel deemed clean on 
arrival but needing to ensure it  is clean before arriving at the next destination. One instance 
of this method has been documented as a pre- departure treatment (Brock et al. 1999).  
While it achieved a significant reduction in the biofouling biomass, it failed to stop the 
introduction of a new species at its destinati on (Apte et al. 2000).  Factors likely to have 
reduced success include the wate r not being entirely fresh (sa linity 6-10 ppt) and the short 
exposure period (9 days).  Freshwater may be a useful treatment option prior to departure if the exposure salinity and time are appropriate, however, it should be noted that many fouling species have wide environmental tolerances so an extended period may be 
required.  The efficacy of freshwater treatment could be enhanced if used in combination 
with other methods, such as scraping to reduc e the bulk of the hull fouling. Freshwater 
exposure has the significant bene fit of being able to reach al l niche areas including internal 
seawater systems.  In New Zealand the use of freshwater as a treatment option, without  encapsulation, faces 
some constraints.  While many New Zeala nd ports are located within estuaries the 
navigable portion of the estuar y is typically salin e (>10 ppt) and often not appropriate for 
larger vessels because of physical (e.g. depth) and/or use (e.g. aquaculture) constraints.  
Westport is currently the only port of first arrival likely to meet the requisite conditions of 
freshwater (<5 ppt) in the port area.  Creating a freshwater lo ck in port environments may 
be possible for some locations, however, th is option may be extremely costly in 
comparison with other hull fouling treatments.   
3.2.3. Land based treatments 
The rem
oval of vessels from the water can be achieved by methods ranging from travel 
lifts to dry docks.  In New Zealand such fac ilities are generally only available for smaller 
vessels with the two major dry dock fac ilities being in Auckland (180 m length, 24 m 
beam) and Lyttelton (130 m length, 14 m beam).    The process of removing a vessel from the wate r can have an associat ed likelihood of entry 
as organisms can be released into the receiving environment during vessel removal. This is especially so when travel- or synchro-lifts or slipways are used. However, unlike in-water 
cleaning, haul outs occur at discrete locations  allowing targeted monitoring and mitigation  
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  29 of possible biosecurity risks.  Moreover, most  facilities are located  near significant ports 
(e.g. Ashby’s shipyard in Opua is only a few hundred metres from the quarantine wharf) 
that are subject to biosecurity surveillance programs.   
 Land based treatments may result in the rel ease and subsequent establishment of non-
indigenous organisms by allowing viable material in the effluent  stream back into the sea 
(McClary & Nelligan 2001, Floerl et al. 2005c, Woods et al. 2006). Although most 
boatyards have coarse gratings to capture large debris and paint chips, this will not stop the 
release of smaller viable propagules such as  gametes. At present there are no MAFBNZ 
certified boatyards, however, it is likely that, in future, boatyards performing waterblasting 
services for biosecurity purposes will have  to comply with th e requirements of the 
MAFBNZ transitional facility standard. For t hose vessels that can be removed from the 
water there are three methods with  which biofouling can be treated. 
3.2.3.1.  Scraping 
Tools ranging from
 a hand scra per to a shovel can be effec tive in rapidly removing large 
scale fouling from a vessel hull.  Scraping is commonly used on ve ssels that require 
fouling removal prior to being placed on a hard stand for maintenance.  The efficacy of this treatment is largely dependant upon the operato r and the types of tools used.  Having a 
vessel out of the water makes it easier for the operator to observe and remove fouling.  
However, removing some types of encrusting or ganisms with a scraper can lead to hull 
damage, such as gouging.  Moreover, scraping is likely to remove only easily visible 
fouling, missing small but viable fragments a nd potentially leaving se ttlement substrates, 
such as remnant barnacle te sts (Floerl et al. 2005b). 
 The key benefits of scraping are that, follo wing haulout and inspec tion, scraping can be 
completed rapidly with tools which are readily accessible
8.  Assuming facilities are 
available, a moderately fouled (<20% of the hull covered in macro-fouling) vessel could be 
returned to the water in less than an hour.    Removing fouling on land has been shown to re duce the proportion of viable organisms in 
the debris relative to in-water cleaning (F loerl et al. 2005c) due to the influences of 
desiccation and trampling.  However, there is a latent risk associated with viable organisms 
in the debris, therefore, it must be ensured that the debris does not return  to the water.  This 
is a requirement for any facility complyi ng with the MAFBNZ transitional facility 
standard.  
3.2.3.2.  Waterblasting 
W
aterblasting is a common tec hnique for fouling removal prio r to further maintenance.  
Waterblasting entails spraying water under pressure (2,000 psi or greater) from a lance, 
often with the water exiting in a triangular, scraper-like pattern.  This system enables the 
operator to systematically clear the hull of fouling with less physical effort, more likelihood of removing microscopic fouling, and less likelihood of damaging the hull. 
However, some anti-fouling systems can be adversely affected by waterblasting.  The 
efficacy of this treatment is dependant  upon the diligence of the operator.   
 For recreational vessels it is possible to dete rmine the time and cost associated with 
standard waterblasting.  Ashby’s boatyard in Opua, for example, advertise a service for 
                                                
 
8 Compared to other mitigation options such as rotating brushes, encapsulation, and heat treatment, which are 
still very much in the process of development and requir e specialised equipement, all that scraping requires is 
often a common paint scraper.  
removing a vessel from the water on the travel lift, waterblasting the hull and returning it 
to the water for less than $600.00 for a 20.3m vessel ( http://www.ashbyboats.co.nz  
accessed 10/03/2009).  However, the costs will  be substantially greater for vessels 
requiring more specialised haul out facilities or dry dock space.  To ensure waterblasting does not result in the release of organisms into the environment, 
the debris removed from the hull and the liquid effluent need to be contained and filtered.   
3.2.3.3.  Desiccation 
Desiccation, or air-drying, is a technique wher e the vessel is hauled out and left on a hard 
stand until all fouling is rendered non-viable.  Although de siccation is m
ost commonly 
used to treat fouling on aquaculture and fi shing equipment, this technique has been 
recognised as having potential for controlling ve ssel biofouling in the biosecurity context 
(Forrest et al. 2004, Coutts & Forrest 2005, Hilli ard et al. 2006).  The benefits of this 
technique are generally consider ed to be that it is easy to perform and does not require the 
use of any environmentally damaging additives.   
 The drawbacks of dessication, however, are cons iderable.  Hilliard et  al. (2006) found that 
“there is no set time which can be considered reliable for applying to all parts of a hull” 
and thus recommended that the minimum period be 21 days, with any reduction requiring expert analysis of the fouling on a case-by-cas e basis.  Given the time and conditions likely 
to be required for effective dessication it is unlik ely that it will be as cost effective as other 
management options.  In addition, the vessel location du ring dessication would need to  be away from any sources 
of moisture, including rainfall,  sea spray and high humidity. Th ese criteria are difficult to 
achieve at most hard stand facili ties.  Moreover, there will be issues of odour as the fouling 
rots and this would need to be mana ged to comply with local regulations. 
 Determining organism viability so a vessel can be declared risk free is also difficult.  
Although a fouling organism may not survive the desiccation process, it could still release 
viable propagules once returned to the water, e.g. Undaria pinnatifida  (Saito 1975).  In 
addition, the presence of dead material on the hull provides an excellent settlement surface 
for biofouling reinfest ation of the hull. 
3.3. REFERENCES 
ASA (2006) Assessment of introduced ma rine pest risks associat ed with niche areas in 
commercial shipping . Australian Shipowners Asso ciation, Melbourne, Australia. 
Aquenal (2008) IMProtector: A quarantine and treatment system fo r apprehended illegal 
foreign fishing vessels . Aquenal Pty Ltd; Tasmania, Australia. 
Bax, N; Hayes, K; Marshall, A; Parry, D;  Thresher, R (2002) Man-made marinas as 
sheltered islands for alien marine organisms: Establishment and eradication of an 
alien invasive marine species. In  Veitch, C R; Clout, M N (Eds) Turning the Tide: 
The Eradication of Invasive Species . IUCN SSC Invasive Sp ecies Specialist Group; 
Auckland, New Zealand; pp 26-39. 
Bohlander, J (2009) Review of options for in-water cleaning of ships . MAF Biosecurity 
New Zealand Technical Paper No: 2009/42. http://www.biosecurity.govt .nz/files/pests/salt-
freshwater/options-for-in- water-cleaning-of-ships.pdf  
30  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
Brock, R; Bailey-Brock, J H; Goody, J (1999)  A case study of efficacy of freshwater 
immersion in controlling introduction of alien marine fouling communities: the USS 
Missouri.  Pacific Science  53: 223-231. 
Callow, M E; Callow, J A (2002) Mari ne biofouling: a sticky problem. Biologist  49(1): 1-
5.  
Coutts, A D M (2002) A biosecurity investigation of a barge in the Marlborough Sounds . 
Cawthron Report No 744. Cawthron Institute; Nelson, New Zealand. 
Coutts, A D M; Forrest, B M ( 2007) Development and applica tion of tools for incursion 
response: lessons learned from the management of a potential marine pest.  Journal of Experimental Mari ne Biology and Ecology  342: 154-162. 
Creese, R G; Davis, A R; Glasby, T M (2004) Eradicating and preventing the spread of 
the invasive alga Caulerpa taxifolia  in NSW .  NSW Fisheries Final Report No. 64, 
NSW Fisheries; Cronulla, Ne w South Wales, Australia. 
Davidson, I C; McCann, L D; Sytsma, M D; Ru iz, G M (2008) Interrupting a multi-species 
bioinvasion vector: The efficacy of in-w ater cleaning for removing biofouling on 
obsolete vessels. Marine Pollution Bulletin  56: 1538-1544. 
Denny, C M (2007) In-situ plastic encapsulation of the NZHMS “Canterbury” frigate: a 
trial of a response tool for marine fouling pests .  Cawthron Report No 1271. 
Cawthron Institute; Nelson, New Zealand. 
Ferguson, R (2000) The effectiveness of Au stralia's response to th e black striped mussel 
incursion in Darwin, Australia . A report of the marine pest incursion management 
workshop, Darwin.  
Floerl, O; Inglis, G J; Hayden, B (2005a) A risk -based predictive tool to prevent accidental 
introductions of non-indigenous marine species. Environmental Management  35: 
765-778. 
Floerl, O; Inglis, G J; Marsh, H M (2005b) Selectivity in vector management: an 
investigation of the effec tiveness of measures used to prevent transport of non-
indigenous species. Biological Invasions  7: 459-475. 
Floerl, O; Norton, N; Inglis, G; Hayden, B; Mi ddleton, C; Smith, M; Alcock, N; Fitridge, I 
(2005c) Efficacy of hull cle aning operations in containing biological material I. Risk 
assessment .  MAF Biosecurity New Zeal and Technical Paper 08/12, MAF 
Biosecurity New Zealand Wellington. 
Forrest, B M; Blakemore, K A (2006) Evaluatio n of treatments to reduce the spread of a 
marine plant pest with aquaculture transfers.  Aquaculture  257: 333-345. 
GISP (2008) Marine Biofouling: An assessment of risks and management initiatives .  
Global Invasive Species Programme a nd the UNEP Regional Seas Programme.  
Godwin, L (2004) Development of an initia l framework for the management of hull 
fouling as a marine invasive species tr ansport mechanism.  Proceedings of a 
workshop on current issues and potential management strategies. Feb. 12 – 13, Honolulu, Hawai’i. 
Hay, C H (1990) The dispersal of sporophytes of Undaria pinnatifida  by coastal shipping 
in New Zealand, and implications for further dispersal of Undaria  in France.  British 
Phycological Journal  25: 301-313 
Hewitt, C L; Ca 
mpbell, M L (2007) Mechanisms  for the prevention of marine bioinvasions 
for better biosecurity.  Marine Pollution Bulletin  55: 395–401 
Hilliard, R; Polglaze, J; LeProvost, I (2006) Review and evaluation of the biofouling 
protocol for vessels less than 25m in length.   Australian Quarantine and Inspection 
Service, Canberra 
Holm, E R; Haslbeck, E G; Horinek, A A (2003) Evaluation of brushes for removal of 
fouling from fouling- release surfaces, using a hydraulic cleaning device.  Biofouling  
19: 297–305. 
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  31  
Hopkins, G A; Forrest, B M ( 2008) Management options fo r vessel hull fouling: an 
overview of risks posed by in-water cleaning.  ICES Journal of Marine Science  65: 
811-815. 
Hopkins, G A; Forest, B M; Coutts, A (In Press) Determining the efficacy of incursion 
response tools: rotating brush technology (coupled with suction capability) .  
Biosecurity New Zealand Technical Pa per, MAF Biosecurity New Zealand, 
Wellington, New Zealand 
Inglis, G; Ponder-Sutton, A;  Unwin, M; Floerl, O (2008) Vessel biofouling as a vector for 
the introduction of non-indigenous marine species to New Zealand: Management 
Tools.  Final report for MAF Biosecurity  New Zealand Research Project 
ZBS2004/04. MAF Biosecurity New Zealand, Wellington 
Inglis, G J; Floerl, O; Unwi n, M; Ponder-Sutton, A; Seawar d, K; Kospartov, M; Bell, A; 
Kluza, D (in prep.). The biosecurity risks associated with biofouling on international 
vessels arriving in New Zealand. Summary of the patterns and predictors of fouling.  
Biosecurity New Zealand Technical Paper. 
Kalumuck, K M; Chahine, G L; Frederick, G S; Aley, P D (1997) Development of a 
DynaJet™ cavitating water jet cleaning tool for underwater marine fouling removal.  
9th American Waterjet Conference 
Lewis, J A (2004) Hull fouling as a vector for the tr anslocation of marine organisms: 
Report 1: Hull fouling research.   Ballast Water Research Series Report No. 14 
Department of Agriculture, Fisheries and Forestry, Canberra 
MAFBNZ (2009) Perna perna  http://www.biosecurity.govt .nz/pests/perna-perna  Accessed 
20 April 2009. 
McClary, D; Nelligan, R (2001) Alternative bi osecurity management tools for vector 
threats: technical guidelines for acceptab le hull cleaning facilities. Ministry of 
Fisheries Project Report ZBS2000-03, Ministry of Fisheries, Wellington, New 
Zealand. 
McEnnulty, F R; Bax, N J; Schaffe lke, B; Campbell, M L (2001) A review of rapid 
response options for the control of ABWMA C listed introduced ma rine pest species 
and related taxa in Australian waters .  Centre for Research on Introduced Marine 
Pests, CSIRO Marine Research, Hobart. 
Piola, R; Johnston, E (2008) The potential for translocation of marine species via small-
scale disruptions to anti-fouling surfaces. Biofouling  24: 145-155. 
Rajagopal, S; Azariah, J; Nair, K V K (1994) Heat treatment as a fouling control method 
for Indian coastal power plants. In Thom pson, M F; Nagabhushanam, R; Sarojini, R; 
Fingerman, M (Eds) Recent developments in biofouling control.  Oxford and IBH; 
New Delhi, India; pp 391-396. 
Redfield, A C; Ketchum, B H (1952) Ship resistance. In Redfield, A C; Ketchum, B H 
(Eds) Marine Fouling and Its Prevention. U.S. Naval Institute; Annapolis, USA. 
Reed, R H; Moffat, L (1983) Copper toxicity and copper tolerance in Enteromorpha 
compressa  (L.) Grev. Journal of Experimental Marine Biology and Ecology  69: 85-
103. 
Russell, G; Morris, O P (1970) Copper tolerance in the marine fouling alga Ectocarpus  
siliceous. Nature  228: 288-289. 
Saito, Y (1975) Undaria . In Tokida, J; Hirose, H (Eds) Advance of phycology in Japan . Dr. 
W. Junk b.v.; The Hague; pp.304-320; 
Statistics New Zealand (2006) New Zealand’s marine economy 1997 – 2002 . Statistics 
New Zealand Environmental Series. St atistics New Zeal and, Wellington. 
Ten Hallers-Tjabbes, C C (2007) Environmentally sound and safe  removal of harmful anti-
fouling systems and of cleaning of ships’ hulls.  SAFEMED Project Task 3.8. Final 
32  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  33 Report prepared under the Project EURO MED Cooperation on Maritime Safety and 
Prevention of Pollution from Ships SAFEMED 
Wallentinus, I (1999) Exotics across the ocean - Undaria pinnatifida  (Harvey) Suringar, 
1872. In Gollasch, S; Minchin, D; Rosenthal, H; Voigt, M (Eds) Exotics across the 
ocean: Case histories on  introduced species: their biology, distribution, range 
expansion and impact .  Institute for Marine Science, University of Kiel, Germany 
pp.11-19 
Woods, C; Floerl, O; Fitridge, I; Johnston, O; Robinson, K; Rupp, D; Davey, N; Rush, N; 
Smith, M (2007) Efficacy of hull cleaning operati ons in containing biological 
material. II. Seasonal variability .  MAF Biosecurity New Zealand Technical Paper 
Series 08/11. MAF Biosecurity New Zealand, Wellington 
Wotton, D M; Hewitt, C L (2004) Marine bios ecurity post-border management: developing 
incursion response systems for New Zealand.  New Zealand Journal of Marine and 
Freshwater Research  38: 553-559. 
Wotton, D M; O'Brien, C; Stua rt, M D; Fergus, D J (2004)  Eradication success down 
under: heat treatment of a sunken tr awler to kill the invasive seaweed Undaria 
pinnatifida . Marine Pollution Bulletin  49: 844-849. 
  
34  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis 4. Risk assessment  
4.1. AMPHIPODS AND ISOPODS  (ORDER: AMPHIPODA AND ISOPODA) 
4.1.1. Hazard Identification 
Am
phipods and isopods are typically small (1- 28 mm long) crustaceans that are widely 
distributed, diverse, abundant, and important components of aquatic ecosystems (Ruppert et 
al. 2004). The order Amphipoda covers over 6,000 species while the order Isopoda has 
approximately 4,000 species (Ruppert et al. 2004). Although most species inhabit marine 
environments, amphipods and isopods are also found in estuarine and freshwater habitats. In 
addition, both orders have terrestrial speci es.  The morphological distinction between 
amphipods and isopods is that amphipods are comp ressed laterally, giving them a shrimp-like 
appearance, whilst isopods ar e dorsoventrally flattened. 
 
The amphipod and isopod life-cycle typically invo lves a short life-span (~1 year), with the 
females spawning multiple times (Ruppert et  al. 2004, Buschbaum & Gutow 2005). Members 
of both orders brood their eggs, clutch sizes  ranging from 2–750 eggs, from which the young 
hatch as miniature adults (R uppert et al. 2004).  These orga nisms do not disperse widely 
outside of their native ranges and their introduction to new environments has been linked to 
either hull fouling or rafting9 (Talley et al. 2001, Buschbaum  & Gutow 2005, Davidson et al. 
2008, Montelli & Lewis 2008).  Ecologically, amphipods and isopods play an impor tant role because they often exhibit high 
densities and diversity (R uppert et al. 2004, Davidson et  al. 2008). Though generally 
omnivorous detritus feeders or s cavengers, both orders contain pr edatory and parasitic species 
(Ruppert et al. 2004, Montelli & Le wis 2008). As predators and mi cro-grazers, some species 
can influence the size distribution, composition, and population dynamics of other epifaunal, 
planktonic, and algal assemblages (Hooff & Bo llens 2004, Marion et al. 2008, Orav-Kotta et 
al. 2009). Amphipods and isopods are also an important food source for larger animals 
(Marion et al. 2008). These orga nisms commonly live in association with habitat forming 
fouling organisms, such as algae, barnacles , and bryozoans (Ruppert  et al. 2004, Montelli & 
Lewis 2008, Guerra-Garcia & Tierno de Figueroa 2009).    Non-indigenous species of amphi pods and isopods, for example, Caprella mutica , Gammarus 
tigrinus, Sphaeroma quoianum  and S. walkeri,  are known to modify habitat, alter food webs, 
and regulate community structure (Cook et al . 2007, Marion et al. 2008, Packalen et al. 2008, 
Piscart et al. 2008, Orav-Kotta et al. 2009). Some isopods, such as members of the family 
Limnoriidae, can cause severe impacts by thei r boring behaviour and thus affect marine 
plants, lateral banks and woode n structures (Borges et al. 2008, Brearley et al. 2008, Davidson 
et al. 2008). By contrast, am phipods and isopods have been observed to occur in high 
densities and abundances with no im pacts identified. For example, S. walkeri  is noted to occur 
in high densities on aquaculture cages with no apparent impacts to the cultured organisms 
(Mak et al. 1985).  Uncertainties regarding the risk presented by amphipods and isopods include a taxonomy that 
is currently poorly known, and limited records of  distributions, biology or ecology (Hewitt et 
al. 1999, Chapman & Carlton 1991, Chapman & Ca rlton 1994, Bollens et al. 2002, Montelli 
                                                
 
9 Organisms or eggs can attach to floating debris (organic or inorganic) and be carried to new lovcations by 
currents.  
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  35 & Lewis 2008). As a result, it is difficult to  determine the extent of amphipod and isopod 
introductions, impacts, local resilience, and importance to ecosystem structure and function 
beyond the few studied species.  
 Given that amphipods and isopods: 
 are known components of marine  biofouling assemblages;  
 are known to have been introduced to new locations; 
 have known impacts on core values; 
it is concluded that amphipod and isopod species  represent a potential hazard via vessel 
biofouling.  
4.1.2. Risk Assessment 
4.1.2.1.  Entry assessment 
Am
phipods and isopods are common and widely distributed in soft and hard substrate 
communities and are reported to occupy sandy beach, rocky shore, continental shelf and 
continental slope environments (Ruppert et al . 2004, Montelli & Lewis 2008).  In the coastal 
environment they are often associated with the natural and artificial substrates of ports and 
marinas. Their mobile nature and swimming ability enables them to move from benthic 
environments onto vessel hulls.  
 As components of hull fouling assemblages,  amphipods and isopods ar e often found living on 
or amongst other hull fouling species which provide food and/or shelter (Buschbaum & 
Gutow 2005, Martinez & Adarraga 2008, Montelli & Lewis 2008).  The associations between 
amphipods, isopods and other hull fouling species can be specific.  For example, Montelli & Lewis (2008) found isopods were mostly associated with acorn barnac les while caprellid 
amphipods were mostly associated with fila mentous algae.  However, tube building 
amphipods have been observed to be the onl y macroscopic organisms living on anti-fouling 
paints under specific test conditions (Lewis 2004) .  Similarly, boring isopods are likely to be 
associated with wooden hulls regardless of th e presence of other fouling species.  While no 
longer common, there remains a small but cons istent influx of wooden hulled recreational 
vessels into New Zealand
10. 
 Amphipods and isopods are known to survive tran slocation on the hulls of vessels.  For 
example, Montelli & Lewis (2008) reported that many species have been translocated on the 
hulls of Australian Naval vessels from both te mperate and tropical waters. In the MAFBNZ 
commissioned research on biofouling Inglis et al. (in prep.) found that, of international vessels 
entering New Zealand, 21% had amphipod and 5% had isopod fou ling the hull, respectively.  
Of those species identified, five amphipods a nd six isopods were not known to be established 
in New Zealand. These examples show that some hull fouling amphipods and isopods can survive transport from international locations.  Given that particular amphipod and isopod species: 
 are known to occur along vessel pathways; 
 are known to be associated with vessel biofouling; 
 are known to survive translocation; 
it is concluded that the likelihood of entr y of non-indigenous amphipod and isopod species 
into New Zealand via vessel biofouling is non-negligible. 
                                                
 
10 11 recreational vessels with wooden hulls, from a samp ling total of 182 recreational vessels, were recorded 
over two years by Inglis et al. (in prep.).   
 
4.1.2.2.  Establishment assessment 
As a
mphipods and isopods are capable swimme rs they could, given the right stimulus, 
migrate from a vessel hull to the receiving enviro nment.  As some species can have high site 
fidelity, the stimulus for disembarking may need to be quite strong, for example, the physical 
dislodgement of associated fouling.  Howe ver, amphipods and isopods, in general, are 
indiscriminate regarding the s ubstratum and fouling community they inhabit.  As a result, 
establishment of these organisms is more likely to be influenced by physical conditions (with 
temperature a known limitation) than substrat e type (Ingolfsson & Agnarsson 2003, Marion et 
al. 2008, Montelli & Lewis 2008).    Port environments are usually sheltered loca tions with high detritus  content, providing a 
suitable habitat for many species of amphipod and isopod. As observed with the establishment of C. mutica  in Europe, ports and artificial hard substrates may act as initial points for 
founding populations (Buschbaum & Gutow 2005). Af ter establishment, it is apparent that 
amphipod and isopod species can subsequently spread throughout th e local estuary or 
immediate coastal area (Cook et al. 2007, Davidson et al. 2008, Marion et al. 2008)    The establishment of certain  species may depend upon the fouling assemblages already 
present due to their specific a ssociations with a particular fo rm or species of macro-fouling 
e.g. caprellid amphipods and filamentous algae (Montelli & Lewis 2008).  However, 
appropriate habitats, to suppor t newly arrived individuals, ar e likely to be common in New 
Zealand port environments.     Once relocated, the ability of amphipods and isop ods to establish sustaining populations is 
supported by their life history characteristics. Specifically, the ability to spawn multiple times 
with each event potentially producing hundreds of  offspring (Ruppert et al. 2004, Cook et al. 
2007).  For example, the invasion success of G. tigrinus  may be dependent on the higher 
reproductive capacity, with regard to wider temp erature and salinity ranges,  compared to the 
native species (Pinkster et al. 1977 In Grabowski et al 2007, Savage 1982 In Grabowski et al 
2007). In addition, G. tigrinus has the ability to thrive in h eavily polluted waters relative to 
native species (Pinkster et al. 1977  In Grabowski et al 2007). 
 Known introductions of amphipods and isopods have  occurred in Australia (Poore & Storey 
1999) and New Zealand (Kospartov et al. in press). Of the 17 non-indigenous amphipods and 
10 non-indigenous isopods recorded in New Zealand, 11 amphipods and two isopods are 
established and all are thought to have been vectored by vessel bi ofouling (Kospartov et al. in 
press).  Given that particular amphipod and isopod species: 
 are known to inhabit similar climates/environmen ts to those that exist in New Zealand; 
 are capable of forming establishi ng populations in New Zealand; 
 have been recently recorded as  established in New Zealand; 
it is concluded that the likelihood of esta blishment of non-indigenous amphipod and isopod 
species in New Zealand via vesse l biofouling is non-negligible. 
 
4.1.2.3.  Consequence assessment 
The consequences of amphipods and isopods es tablishing in New Zeal and could be wide 
ranging. High abundances of a
mphipod and isopo d species have been shown to affect 
zooplankton, benthic habitats and communities. For example, G. tigrinis,  introduced to 
36  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
Finland, has outcompeted and replaced native herbivorous Gammarus  species (Packalen et al. 
2008). Similarly, G. tigrinus has the potential to outcompete the native G. salinus  within 
Pilayella littoralis dominated macrophyte beds as, in laboratory experiments, the presence of 
G. tigrinus  within Pilayella littoralis  induced a higher mortality of G. salinus (Orav-Kotta et 
al. 2009). Therefore, the recent blooms of P. littoralis  are expected to favour the formation of 
G. tigrinus  dominated communities in the coastal range of the northern Baltic Sea. The 
removal of native amphipods and isopods as inte rmediaries in the food chain may have flow 
on effects for the organisms dependant upon them. 
 Boring isopods have relatively well known impact s on marine structures and habitat. For 
example, S. terebrans  bores into mangroves and S. quoianum, a wood borer in its native 
Australian range, bores into banks in North Amer ica, accelerating erosion in estuaries (Talley 
et al. 2001, Davidson et al. 2008). Talley et al. (2001) observed a correlation between the rate 
of lateral bank loss and density of Sphaeroma spp. burrows, whereby Sphaeroma spp. could 
increase the rate of sediment loss by as much as 240%. In some areas where Sphaeroma spp. 
is abundant, lateral erosion exceed s one metre per year (Talley et  al. 2001). Erosion of banks 
may increase the sedimentation of estuaries lead ing to light attenuation, reduced water quality 
and the smothering of benthic organisms.  The taxonomy of amphipods and isopods is poorly known, and there are few records of 
distributions, biology or eco logy (Chapman & Carlton 1991, Chapman & Carlton 1994, 
Hewitt et al. 1999, Bollens et al. 2002, Montelli & Lewis 2008). As a consequence, it is 
difficult to determine the extent of am phipod and isopod introductions, impacts, local 
resilience, and importance to  ecosystem structure and f unction beyond the few studied 
species.   
4.1.3. Risk Estimation 
Given that particular amphipod and isopod species: 
 have the po
tential to enter New Zealand via vessel biofouling; 
 are capable of forming establishi ng populations in New Zealand; 
 have recently established populations within New Zealand; 
 are known for their invasive behaviours; 
 have known recorded imp acts to core values; 
it is concluded that the import of non-indi genous amphipod and isopod species via biofouling 
poses a non-negligible risk to Ne w Zealand’s marine core values. 
4.1.4. Risk Management 
A generic approach is suitable to m
anaging th e risk associated with amphipods and isopods 
(Section 3). However, the mobility of amphipods and isopods presents further management 
difficulties, and as a result, systems that contai n all organisms in-water prior to treatment may 
be required to reduce the risk of pre-treatment escape. The fact that amphipods can also form tubes on vessel hulls devoid of macro-f ouling (Montelli & Lewis 2008) should also be 
considered during inspections.   
4.1.5. References 
Berezina, N A (2007) Invasions of alien am phipods (Am 
phipoda: Gammaridea) in aquatic 
ecosystems of North-Western Russ ia: pathways and consequences. Hydrobiologia 590: 
15-29. 
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  37  
Bollache, L; Dick, J T A; Farnsworth, K D;  Montgomery, W I (2008) Comparison of the 
functional responses of inva sive and native amphipods. Biology Letters  4: 166-169. 
Bollens, S M; Cordell, J R; Avent, S; Hooff,  R (2002) Zooplankton inva sions: a brief review, 
plus two case studies from the northeast Pacific Ocean. Hydrobiologia 480: 87-110. 
Borges, L M S; Cragg, S M; Bergot, J; Willia ms, J R; Shayler, B; Sawyer, G S (2008) 
Laboratory screening of tropica l hardwoods for natural resistance to the marine borer 
Limnoria quadripunctat a: the role of leachable and non-leachable factors. 
Holzforschung  62: 99-111. 
Brearley, A; Kendrick, G A; Walker, D I (2008) How does burrowing by the isopod Limnoria 
agrostisa  (Crustacea: Limnoriidae ) affect the leaf canopy of southern Australian 
seagrass Amphibolus graffithii ? Marine Biology 156: 65-77. 
Buschbaum, C; Gutow, L (2008) Mass occurrence of an introduced crustacean ( Caprella cf. 
mutica ) in the south-eastern North Sea. Helgoland Marine Research  59: 252-253. 
Chapman, J W; Carlton, J T (1991)  A test of criteria for intr oduced species: the global 
invasion by the isopod Synodotia laevidorsalis  (Miers, 1881). Journal of Crustacean 
Biology 11: 386-400. 
Chapman, J W; Carlton, J T (1994) Predic ted discoveries of the introduced isopod Synidotea 
laevidorslis  (Miers, 1881).  Journal of Crustacean Biology  14: 700-714. 
Cook, E J; Jahnke, M; Kerckhof, F; Minchin, D; Faasse, M; Boos, K; Ashton, G (2007) 
European expansion of the introduced amphipod Caprella mutica  Schurin 1935. 
Aquatic Invasion  2: 411-421. 
Cordell, J R; Morrison, S M (1996)  The invasive Asian copepod Pseudodiaptomus inopinus  
in Oregon, Washington, and Brit ish Columbia estuaries.  Estuaries 19: 629-638. 
Cordell, J R; Rasmussen, M; Bollens, S M (2007) Biology of the introduced copepod 
Pseudodiaptomus inopinus  in a northeast Pacific estuary. Marine Ecology Progress 
Series  333: 213-227. 
Davidson, T M; Hewitt, C L; Campbell, M (2008 ) Density, distribution and habitat use among 
native and introduced populations of the Australasian burrowing isopod Sphaeroma 
quoianum .  Biological Invasions  10: 399-410. 
Guerra-Garcia, J M; Tierno de Figueroa, J M (2009) What do caprellids (Crustacea: 
Amphipoda) feed on? Marine Biology 156: 1881-1890. 
Gewant, D S; Bollens, S M (2005) Macrozooplankton and mi cronekton of the lower San 
Francisco estuary: seasonal, interannual,  and regional variation in relation to 
environmental conditions.  Estuaries  28: 473-485. 
Grabowski, M; Ja żdżewski, K; Konopacka, A (2007) Alien Crustacea in Polish waters – 
Am
phipoda. Aquatic Invasions  2(1): 25-38. 
Hewitt, C L; Campbell, M L; Thresher, R E; Martin, R B (1999) Marine biological invasions 
of Port Phillip Bay, Victoria .  Centre for Research on Introduced Marine Pests 
Technical Report No. 20. CSIRO Mari ne Research; Hobart, Australia. 
Inglis, G J; Floerl, O; Unwi n, M; Ponder-Sutton, A; Seawar d, K; Kospartov, M; Bell, A; 
Kluza, D (in prep.). The biosecurity risks associated with biofouling on international 
vessels arriving in New Zealand. Summary of the patterns and predictors of fouling.  
Biosecurity New Zealand Technical Paper. 
Ingolfsson, A; Agnarsson, I (2008)  Amphipods and isopods in the rocky intertid al: dispersal 
and movements during high tide. Marine Biology 143: 859-866. 
Kelly, D W; Muirhead, J R; Heat h, D D; MacIsaac, H J (2006) C ontrasting pattern s in genetic 
diversity following multiple invasions of fresh and brackish waters. Molecular Ecology  
15: 3641-3653.  
Kospartov K; Inglis G; Seawar d K; Brink A van den; D’Arch ino R; Ahyong S (in press) Non-
indigenous and cryptogenic mari ne species in New Zealand . Current state of knowledge 
– Interim report prepared for MAFBNZ Post Border Directorate. MAF Biosecurity New 
Zealand. Wellington, New Zealand. 
38  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  39 Lewis, J (2001) Hull fouling as a vector for th e translocation of marine  organisms. Report 1: 
Hull Fouling Research.  Ballast Water Research Series Report  No. 14.  Department of 
Agriculture, Fisheries and Forestry; Canberra, Australia. 
NIMPIS (2010) Sphaeroma walkeri  species summary. In Hewitt, C L; Martin, R B; Sliwa, C; 
McEnnulty, F R; Murphy, N E; Jones, T; C ooper, S (Eds) National Introduced Marine 
Pest Information System. http://crimp.marine.csiro.au/nimpis - accessed 17 June 2008.  
Mak, P M S; Huang, Z G; Morton, B S (1985) Sphaeroma walkeri  Stebbing (Isopoda, 
Sphaeromatidae) introduced into  and established in Hong Kong. Crustaceana 49: 75-
82. 
Marion, A; Harvey, M (2008) Feeding ecology and predation impact of the recently 
established amphipod, Themisto libellula , in the St. Lawrence marine system, Canada. 
Marine Ecology Progress Series  373: 53-70. 
Martinez, J; Adarraga, I (2008) Firs t record of invasive caprellid Caprella scuara Templeton, 
1836 sensu lato (Crustacea: Amphipoda: Caprellidae) from the Iberian Peninsula. 
Aquatic Invasions  3: 165-171. 
Montelli, L; Lewis, J (2008) Survey of Biofouling on Australian Navy Ships: Crustacea; 
Isopoda and Amphipoda; Caprellidea. DSTO Defence Science and Technology 
Organisation; Victoria, Australia. 
Morton, J (2004) Seashore Ecology of New Zealand and the Pacific. David Bateman Ltd; 
Auckland, New Zealand. 
Orav-Kotta, H; Kotta, J; Herkul , K; Kotta, I; Paalme, T (2009)  Seasonal variability in the 
grazing potential of the invasive amphipod Gammarus tigrinus  and the native amphipod 
Gammarus salinus  (Amphipoda: Crustacea) in the northern Baltic Sea. Biological 
Invasions  11: 597-608. 
Packalen, A; Korpinen, S; Lehtonen, K K (2008) The invasive amphipod species Gammarus 
tigrinus (Sexton 1939) can rapidly change littoral communities in the Gulf of Finland 
(Baltic Sea). Aquatic Invasions  3: 405-412. 
Poore, G C B; Storey, M (1999) Soft sediment crustacea of Port  Phillip Bay. In, Hewitt, C L; 
Campbell, M L; Thresher, R E; Martin, R B (Eds) Marine Biological Invasions of Port 
Phillip Bay, Victoria . Centre for Research on Introduced  Marine Pests Technical Report 
20. CSIRO, Marine Research, Hobart, Australia. 
Piscart, C; Maazouzi, C; Marmonier, P (2008) Range expansion of the North American alien 
amphipod Gammarus tigrinus  Sexton 1939 (Crustacea: Gammaridae) in Brittany, 
France. Aquatic Invasion s 3: 449-453. 
Ruppert, E E; Fox, R S; Barnes, R D (2004) Invertebrate Zoology: A functional evolutionary 
approach (7th edition). Thompson Lear ning; Victoria, Australia. 
Talley, T S; Crooks, J A; Levin, L A (2001) Habitat utilization and alteration by the invasive 
burrowing isopod, Sphaeroma quoyanum , in California salt marshes. Marine Biology 
138: 561-573. 
  
4.2. BACTERIA AND VIRUSES 
4.2.1. Hazard identification 
Bacteria are well known com
pone nts of the biofilms which rapidly develop on submerged 
surfaces (Skerman 1956, ZoBell 1959, Wahl 1989, Cooksey & Wigglesworth-Cooksey 1995, 
Callow & Callow 2002).  Bacteria in biofilms may be  a single species or a complex of species 
and are known to function cooperatively providing the community  with a level of protection 
against predation, stressors (suc h as salinity, temperature or ultraviolet shock) and sheer 
forces (Davey & O’Toole 2000).  Moreover, bi ofilms have been observed to modify their 
structure to conform to changi ng external stressors such as current flows (Purevdorj et al. 
2002, Hall-Stoodley et al. 2004).    While in the past there has been a view that  marine bacterial species are omnipresent, 
improved identification techniques have begun to show a distinct bacterial biogeography 
(Martiny et al. 2006, Ramette & Tiedje 2007).  Furhman et al. (2008), for example, have 
reported “geographic patterns of [marine bacteria l] species diversity that are similar to those 
seen in other organisms” noting a strong correlation between species richness, latitude and temperature.  Although such observations are re cent, they suggest that bacteria could be 
translocated, via anthropogenic vectors, beyond their natural limits with similar consequences to other invasive organisms.  Indeed this ha s been observed in the context of ballast water 
translocation (Drake et al. 2001). However, de spite its volume, balla st water distributes 
primarily pelagic bacterial species whereas the greatest numbers and dive rsity of bacteria are 
those attached to surfaces as biofilm (Davey & O’Toole 2000).  Potentially pathogenic species f ound in marine biofilms include Vibrio parahaemolyticus,  
Pseudomonas aeruginosa and P. putrefaciens  (Zambon et al. 1984, Lee et al. 2003, Railkin 
2004).  However, research on the biofilms associat ed with vessel hulls has tended to focus on 
its association with biofouling macro-species (Cassé & Swain 2006).  Similarly, few investigations of biofilms on natural substrates  in the marine environment have considered 
pathogens.  However, it has been confirmed that  pathogens are associated with biofilms and 
that pathogens have the abil ity to exit biofilms to fac ilitate cell distribution (Boyd & 
Chakrabarty 1995, Decho 2000, Hall-S toodley et al. 2004).  Pathoge nic bacteria have been 
isolated from the biofilm of ba llast water tanks (Drake  et al. 2005), similar to the findings of 
more intensively researched systems, e.g. m unicipal drinking water supplies (Walker et al. 
1995).  Virus-like particles have al so been reported from ballast water tank biofilms (Drake et 
al. 2005) and it is likely that they would be associated with hull biofouling.  However, to date, 
there have been no known bacterial or vi ral outbreaks attribut ed to biofouling. 
 The MAFBNZ commissioned research on biofouli ng showed the presence of cyanobacteria in 
some samples of passenger and commercial vessel  biofouling, and no viruses (Inglis et al. in 
prep.).  However, it is noted that the samp ling methods focused on macro-fouling organisms 
rather than micro-organisms.  A knowledge gap needs to be a ddressed regarding the potential 
of pathogenic bacteria and viru ses to be translocated beyond th eir natural geographic ranges.  
Initial investigations of biof ouling related pathogens are cu rrently being conducted by the 
MAFBNZ Wallaceville National Centre for Bi osecurity and Infectious Diseases.   
 Given that bacteria and viruses; 
 are well documented components of biofilms; 
 are likely to be present on all vessel hulls; 
40  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
 of human pathogenic form are known to be pr esent in biofilms though have yet to be 
specifically associated with vessel hull fouling; 
 have an unknown potential to establish and b ecome problematic in New Zealand’s marine 
environment; 
it is concluded that bacteria and virus species represent a poten tial hazard via vessel 
biofouling. 
 As current anti-fouling system t echnologies only have the capacity  to control biofouling to the 
biofilm (slime layer) level, th e prevention of the entry of non- indigenous bacteria and viruses 
via vessel hull fouling cannot be practically achie ved. As a consequence, further research is 
required to confirm the presence and transloca tion potential of potentially hazardous bacteria 
and viruses before the development and implem entation of biofilm control measures can be 
justified specifically from the biosecurity perspective.  
4.2.2. References 
Boyd, A; Chakrabarty, A 
 M (1995) Pseudomonas aeruginosa  biofilms: role of the alginate 
exopolysaccharide.  Journal of Industrial Micr obiology and Biotechnology  15: 162-
168. 
Callow, M E; Callow, J A (2002) Mari ne biofouling: a sticky problem. Biologist  49: 1-5. 
Cassé, F; Swain, G W (2006) The development of micro-fouling on f our commercial anti-
fouling coatings under static and dynamic immersion.  International Biodeterioration 
and Biodegradation  57: 179-185. 
Cooksey, K E; Wigglesworth-Cooksey, B (1995)  Adhesion of bacteria and diatoms to 
surfaces in the sea: a review. Aquatic Microbial Ecology  9: 87-96. 
Davey, M E; O’Toole, G A (2000)  Microbial biofilms: from eco logy to molecular genetics. 
Microbiology and Molecular Biology Reviews  64: 847-867. 
Decho, A W (2000) Microbial biofilms in intertidal systems: an overview.  Continental Shelf 
Research  20: 1257-1273. 
Drake, L A; Choi, K-H; Ruiz, G M; D obbs, F C (2001) Global  redistribution of 
bacterioplankton and virioplankton communities.  Biological Invasions  3: 193-199. 
Drake, L A; Meyer, A E; Forsberg, R L; Baie r, R E; Doblin, M A; Heinemann, S; Johnson, W 
P; Koch, M; Rublee, P A; Dobbs, F C (2005)  Potential invasion of microorganisms 
and pathogens via ‘interior hull fouling’: biofilms inside ballast water tanks.  
Biological Invasions  7: 969-982. 
Fuhrman, J A; Steele, J A; He wson, I; Schwalbach, M S; Brow n, M V; Green, J L; Brown, J 
H (2008) A latitudinal diversity gradie nt in planktonic marine bacteria.  Proceedings 
of the National Academy of Sciences  105: 7774-7778. 
Hall-Stoodley, L; Costerton, J W; Stoodley, P (2 004) Bacterial biofilms: from the natural 
environment to infectious diseases.  Nature Reviews Microbiology 2: 95-108. 
Inglis, G J; Floerl, O; Unwi n, M; Ponder-Sutton, A; Seawar d, K; Kospartov, M; Bell, A; 
Kluza, D (in prep.). The biosecurity risks associated with biofouling on international 
vessels arriving in New Zealand. Summary of the patterns and predictors of fouling.  
Biosecurity New Zealand Technical Paper. 
Martiny, J B H; Bohannan, B J M; Brown, J H; Colwell, R K; Fuhrman, J A; Green, J L; 
Horner-Devine, M C; Kane, M; Krumins, J A; Kuske, C R; Morin, P J; Naeem, S; 
Øvreas, L; Reysenbach, A-L;  Smith, V H; Staley, J T ( 2006) Microbial biogeography: 
putting microorganisms on the map. Nature Reviews Microbiology  4: 102-112. 
Purevdorj, B; Costerton, J W; Stoodley, P (2002) Influence of hydrodynamics and cell 
signalling on the structure and behaviour of Pseudomonas aeruginosa biofilms. 
Applied and Environmental Microbiology  68: 4457-4464. 
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  41  
42  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis Railkin, A I (2004) Marine biofouling: coloni zation processes and defences. CRC Press, Boca 
Raton. 
Rammette, A; Tiedje, J M (2007) Biogeography: An emerging cornerstone for understanding 
prokaryotic diversity, ecology, and evolution. Microbial Ecology 53: 197-207. 
Skerman, T M (1956) The nature and developmen t of primary films on surfaces submerged in 
the sea. New Zealand Journal of Science and Technology  38B: 44-57. 
Wahl, M (1989) Marine epibiosis. I. Fouling and anti-fouling: some basic aspects. Marine 
Ecology Progress Series  58: 175-189. 
Walker, J T; Mackerness, C W; Rogers, J; K eevil, C W (1995) Heteroge neous mosaic biofilm 
– A haven for waterborne pathogens. In Lappin-Scott, H M; Costerton, J W (Eds) 
Microbial Biofilms . Cambridge University Press;  Cambridge, United Kingdom. 
Zambon, J J; Huber, P S; Meyer, A E; Slots, J; Fornalik, M S; Baier, R E (1984) In situ 
identification of bacterial species in marine micro-fouling films by using an 
immunofluorescence technique. Applied and Environmental Microbiology  48: 1214-
1220. 
ZoBell, C E (1959) Introduction to marine microbiology. In Skerman, T M. Contributions to 
marine microbiology. New Zealand Department of Scient ific and Industrial Research 
Information Series  22: 7-20. 
   
4.3. BARNACLES (ORDER: SESSILIA) 
4.3.1. Hazard Identification 
Barnacles are the common nam
e for the group of  arthropods classified  within the class 
Maxillopoda. However, for simplicity, within this  hazard identification and risk assessment 
the term “barnacle” will be used to refer to  the order Sessilia with in the Maxillopoda. The 
order Sessilia is composed of several families; the Verrucidae are a deep water family, while 
the remaining families make up a group commonly referred to as Balanomorpha (Species 2000, Perez-Losada et al. 2004).  Although the or der Pedunculata (goose or stalked barnacles) 
are also commonly known as barnacles these are found most often in th e pelagic environment 
(Barnes et al. 2004, Lewis 2004) and are theref ore not considered in this assessment. 
 Adult epibenthic barnacles live permanently attached to the surf ace of both natural and 
artificial hard substrat es and disperse via a planktonic la rval phase (Skerman 1960, Barnes et 
al. 2004, Lawson et al. 2004). Usually found in conspecific clumps, barnacles reproduce 
sexually and must be located within penis reach to achieve a se lf sustaining population 
(Barnes & Crisp 1956, Kent et al. 2003, Aldred & Clare 2008). Larvae ar e released into the 
plankton and develop over a period from days to weeks before settlement (Callow & Callow 2002). Attachment is head-down and the larval carapace, which encloses the entire body, persists and becomes a shell of calcareous plates. The top plates open to allow the thorax to be 
extended for filter feeding (Miller 2007). The ad hesive used in cementation and exploration 
contains chemical cues that induce the settle ment of other barnacle larvae (Callow & Callow 
2002).  Barnacles inhabit intertidal and shallow coas tal waters and are important as ecosystem 
engineers. The thickly calcified shell and valv es protect against pred ation, desiccation and 
sheer stresses (Foster 1971, Ruppe rt et al. 2004). They grow cem ented directly to a surface 
from which it takes considerable force to remove (Crisp et al. 1985). As a result, barnacles are 
able to live in stressful environments, such as the rocky intertidal, with little competition for 
space and therefore can occur in high densitie s (Ruppert et al. 2004, Mill er 2007). Due to their 
dominance of space and hard structures, barnacl es can provide shelter and attachment points 
for an assortment of other organi sms (Leppakoski & Olenin 2000).  
 Barnacles are well known nuisance foulers of vess el hulls and other ar tificial structures 
(Railkin 2004, Sakaguchi & Namihira 2006). As th e speed and fuel efficiency of a barnacle-
fouled vessel may be reduced by up to 30%, extens ive research has inves tigated their control 
(Railkin 2004, Aldred & Clare 2008). The cemen tation of barnacles to  artificial surfaces 
means that removal can be a difficult and tim e consuming task, often leaving residue or 
damaging the surface of the anti-fouling system (Thiyagarajan et al. 1997, South Carolina Department of Natural Resources 2006). This residue and damage can provide a surface on to which other fouling species can attach. 
 
Introduced globally, non-indigenous species of ba rnacles can modify co mmunity structures 
by excluding native communities of barnacles , mussels, or macroalgae (Kado 2003, Lawson 
et al. 2004, Watson et al. 2005, Gell er et al. 2008). I nvasive barnacles have also covered 
previously bare space creating novel habitat in Hawai’i and eastern South America (Zabin & 
Altieri 2007, Laird & Griffiths 2008). Such co mmunity and habitat modification has also 
affected mobile intertidal species, such as the native Hawai’ian limpet Siphonaria normalis  
(Zabin & Altieri 2007). Non-indigenous barnacles dominate sectio ns of the coast in various 
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  43  
countries, such as Ireland and Japan, and have restricted the habitats  of native barnacles 
(Kado 2003, Lawson et al. 2004, Watson et al. 2004). 
 Given that barnacles: 
 are known components of marine  biofouling assemblages;  
 are known to have been introduced to new locations; 
 have known impacts on core values; 
it is concluded that barnacle species represen t a potential hazard vi a vessel biofouling. 
 
4.3.2. Risk Assessment 
4.3.2.1.  Entry assessment 
Barnacles occur in coastal environments, on both natural and artificial hard substrates, from
 
latitudes 57oS to 70oN (Barnes et al. 2004). Some speci es inhabit port and harbour 
environments, thus placing them in close proxim ity to vessels. Vessel hulls present an ideal 
settlement option for barnacle larvae as they are attracted to open hard surfaces that generally 
free of other biofouling, and to other barnac les (Olivier et al. 2000, Anderson et al. 2003, 
Ruppert et al. 2004). Moreover, the sheltered nature of ports  and harbours retain pelagic 
larvae, thus increasing the likelihood of settleme nt onto vessels (Floerl et al. 2003, Floerl et 
al. 2005, Davidson et al. 2008).  
 Barnacles are highly prevalent hull fouli ng organisms (Wood & Allen 1958, Christie & 
Dalley 1987 In Floerl & Inglis 2003, Aldred & Cl are 2008, Hung et al. 2008). As such, 
biocidal anti-fouling paints have been devel oped to have high efficacy against barnacles by 
altering the attachment characte ristics of metamorphosing larv ae thus causing juveniles to 
detach more easily (Crisp & Austin 1960, Floerl et  al. 2005). However, if a juvenile or adult 
barnacle is removed any basal membrane or cem ent remaining will act as an attractant to 
other settling larvae (Crisp 1961, Crisp & Meadows 1962). In addition, settlement on these remnants provides protection from anti-fouling biocides. Thus improper cleaning of hulls may facilitate the settlement of barnacles by leavi ng remnant tests or small-scale disruptions to the 
anti-fouling system (Floerl et al. 2005, Piola & Johnston 2008).  
Barnacles are habitat structuring organisms due to  their gregarious settling habit.  Thus, as 
hull fouling, barnacles may ai d the recruitment of other f ouling organisms by providing 
protection from sheer forces a nd predation (Railkin et al. 1985, Leppakoski & Olenin 2000, 
Minchin & Gollasch 2003).     
 Barnacles are tolerant to environmental stressors, aiding their survival  during translocation. 
For example, the temperate barnacles Austrominius modestus  and Balanus glandula  have 
been translocated between hemi spheres indicating survival thr ough the thermal barrier of the 
equator (O’Riordon & Ramsay 1999, Lepakoski  & Olenin 2000, Geller et al. 2008, Simon-
Blecher et al. 2008). Similarly the spr ead of tropical barnacles, such as Chthamalus proteus , 
indicates that they can withstand passages thr ough cold or freshwater environments (Zardus & 
Hadfield 2005).    Hull fouling is the primary mechanism by whic h barnacles have been known to enter New 
Zealand (Kospartov et al. in press). In th e MAFBNZ commissioned research on biofouling, 
Inglis et al. (in prep.) found that 60% of sampled vessels en tering New Zealand had barnacle 
fouling. In addition, 42 of the 45 species reco rded were not established in New Zealand 
(Inglis et al. in prep.). Similarly, Otani et al. (2007) found 22 species of barnacles on bulk 
carriers in Osaka Bay, Japan, of which 14 had not  been previously recorded in the bay. 
44  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
Genetic analysis of the wide ly, and recently, distributed Megabalanus coccopoma , M. rosa 
and M. volcano  suggest that shipping was the transloc ation vector (Yama guchi et al. 2009).  
 
Given that particular barnacle species: 
 are known to occur along vessel pathways; 
 are known to be associated with vessel biofouling; 
 are known to survive translocation; 
it is concluded that the likelihood of entry of non-indigenous barnacle species into New 
Zealand via vessel biofou ling is non-negligible. 
 
4.3.2.2.  Establishment assessment 
It is likely that non-indigenous barnacles will find a suitable habitat in N
ew Zealand given 
that approximately half the shoreline is com posed of hard substrata (Carter & Saunders 2003). 
In addition, artificial substrat es, such as provided by ports, yield favourable conditions for the 
establishment of sustaining populations (Sakaguchi & Namihira 2006, Weigemann 2008).  
 Barnacles are most likely to establish in a ne w environment through the release of larvae.  
Barnacles release large numbers of planktonic la rvae which can be widely dispersed prior to 
attaching to a suitable substrate (Jarre tt 1997, Thiyagarajan 2002a , Thiyagarajan 2002b, 
Thiyagarajan 2003). Larval release generally occurs during summer months, however, some 
species, such Elminius modestus , release larvae year-round (Skerman 1960).  
 For a barnacle population to be re productive it must reach critical density, i.e. individuals of 
the same species must be able to reach each other for copulation to occur (Barnes & Crisp 
1956, Kent et al. 2003, Aldred & Clare 2008). As a result, newly settled barnacles must be aggregated to create a sustainable populati on. While barnacles have been known to survive 
dislocation from a hull (Woods et al. 2007), a debris  field with live barnacles is not likely to 
support a reproductive population due to continual disturbance. However, in areas of 
restricted flow, such as sheltered harbours and ports, larvae can be retained increasing the 
number of recruits and, therefore, the like lihood of gregarious and successful settlement 
(Floerl et al. 2003, Floerl et al. 2005). Furthermore, the pres ence of conspecific adult 
barnacles, or their remnant tests, can be an  important settlement cue (Jeffery 2002, NIMPIS 
2010). The ability of barnacles to settle on artifi cial substrates in expos ed areas not tolerated 
by other taxa means they can form beachhead po pulations from which they can be spread. For 
example, Austrominius modestus  inhabits sheltered locations and has been spread around the 
Irish coastline through vessel tr ansport (Allen et al. 2006).  
 Water temperature is a critical factor to th e success of non-indigenous barnacle species as 
development, settlement, and metamorphosis of larvae all depend upon this factor (Watson et al. 2005, Weigemann 2008). For example, although Amphibalanus amphitrite larvae have 
been found in German waters they have not established on the shore because it is too cold for the larvae to fully develop (Wiegemann 2008). Similarly, in Peter the Great Bay, Sea of 
Japan, cold waters not only rest rict the reproductive window of A. amphitrite  but also result in 
winter mortalities (Zvyagintsev & Korn 2003). Barnacles that establish in New Zealand are 
likely to have subtropical or temperate di stributions. For exampl e, in the MAFBNZ 
commissioned research on biofouling, approxi mately half of the non-indigenous barnacle 
species recorded entering New Zealand on recr eational vessels had nati ve distributions in 
temperate environments (Inglis et al. in prep.). Therefore, it is likely that these species will find suitable conditions for establishment. Kosp artov et al. (in press)  found that six non-
indigenous barnacles have recently establis hed in New Zealand. Of those, the two 
Austromegabalanus  species are noted to have established since 2005.  
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  45  
 
Given that particular barnacle species: 
 are known to inhabit similar climates/environmen ts to those that exist in New Zealand; 
 are capable of forming establishi ng populations in New Zealand; 
 have been recently recorded as  established in New Zealand; 
it is concluded that the likelihood of establis hment of non-indigenous barnacle species in New 
Zealand via vessel biofou ling is non-negligible. 
 
4.3.2.3.  Consequence assessment 
The consequ
ences of barnacles are increased main tenance and transport costs, threats to local 
biodiversity, and the facilit ation of the entry of othe r non-indigenous organisms.  
 
The further introduction of non-indigenous hull f ouling barnacles to New Zealand will impact 
maintenance of hulls and fuel efficiency, especia lly given that some are very large, such as,  
Balanus  cf rostatus , Perforatus perforatus , and Austrobalanus imperator . For example, in 
Sweden the economic impact of the bay barnacle B. improvisus , calculated as fouling control 
costs, is estimated at 166-418 million Swedish kr ona per year, of which recreational vessels 
account for approximately 74% (Gren et al . 2007). Although not al l these costs were 
specifically due to B. improvisus , this organism is considered the most problematic fouling 
organism in Swedish waters. The environmen tal costs caused by toxic compounds used in 
anti-fouling paints are not acc ounted for in this cost es timate (Gren et al. 2007). 
 Non-indigenous barnacles occupyi ng natural or artificial substrat a in the intertidal zone may, 
through competition or the presence of an ope n niche, alter the local marine community. 
Impacts may include reduction or elimination of local barnacles, decreased mobility for 
organisms such as limpets, or elimination of organisms such as mussels and algae. For example, the Hawai’ian archipelago was generall y without high cover of intertidal 
barnacles until the recent invasion by Chthamalus proteus. C. proteus  has become highly 
abundant in the low-energy environments of the island of Oahu, whereby, cover of C. proteus 
can extend from the high to low intertidal and reach up to 100% cover in bays on the windward side. Field surveys and experimental re sults suggest that the creation of suboptimal 
grazing conditions by barnacle recru itment has resulted in limpets, Siphonaria normalis, 
moving to patches with lower numbers of barnacles (Zabin  & Altieri 2007). In Argentina, the 
introduced B. glandula,  is presently distributed over 2900 km, with an estimated dispersion 
rate of 244 km per year (Swindt 2007). In most intertidals of Argentina B. glandula  has 
formed dense monospecific belts devoid of competition from other barn acle species (Swindt 
2007).  As habitat providers and common foulers that deposit calcareous  shells, barnacles have the 
potential to facilitate many hull fouling groups. Both mobile and attached organisms, such as 
crabs, macroalgae, ostracods, copepods, and juve nile bivalves have been associated with 
barnacle hull fouling (Railkin  et al. 1985, Leppakoski & Ol enin 2000, Minchin & Gollasch 
2003). The consequences of these introductions ar e discussed within the respective group risk 
assessments.      
4.3.3. Risk Estimation 
Given that particular barnacle species: 
 have been frequently recorded enteri ng New Zealand via vessel fouling; 
 have reco
rds of recent establishment in New Zealand; 
 have known invasive behaviour; 
46  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
 have been introduced globally via vessel fou ling with, in some cases , significant impacts; 
 can be ecosystem engineers; 
 can facilitate the entry and es tablishment of a variety of other non-indigenous organisms; 
 are likely to impact a range New Zeala nd’s core values, such as, economic and 
environmental values; 
it is concluded that the impor t of non-indigenous barnacle species via biofouling poses a non-
negligible risk to New Zeal and’s marine core values. 
 
4.3.4. Risk Management 
The trea
tment of hulls to manage the risks posed by barnacles is covered by the generic 
options (Section 3).  An effective management of  barnacle risk is their removal from vessel 
hulls, as the removal process is likely to result in barnacle mo rtality. By contrast,  application 
of in-situ  techniques, such as encapsulation, are ch allenging due to the high tolerance of 
barnacles to adverse conditions.  The choice of removal technique should consider the 
complete removal of the test, including cement, and the contai nment of potential larval and 
associated fouling releases. 4.3.5. References 
Aldred, N; Clare, A S (2008) The adhesive strategies of cyprid s and developm 
ent of barnacle-
resistant marine coatings. Biofouling  24(5): 351-363. 
Barnes, H; Crisp, D J (1956) Evid ence of self-fertiliza tion in certain species of barnacles. 
Journal of the Marine Biologica l Association of the United Kingdom 35: 631–639. 
Barnes, D K A; Warren, N L;  Webb, K; Phalan, B; Rei d, K (2004) Polar pedunculate 
barnacles piggy-back on pycnogona, pengui ns, pinniped seals and plastics. Marine 
Ecology Progress Series  284: 305-310.  
Bertness, M D; Gaines, S D; Yeh, S M (1998)  Making mountains out of barnacles: the 
dynamics of acorn barnacle hummocking. Ecology  74(4): 1382-1394. 
Callow, M E; Callow, J A (2002) Mari ne biofouling: a sticky problem. Biologist  49(1): 1-5.  
Carter, L; Saunders, H (2003) Rocky Reef s. In Andrew N; Francis M (Ed) The Living Reef – 
the ecology of New Zealand’s rocky reefs . Craig Potton Publis hing; Nelson, New 
Zealand; pp 18-23.  
Crisp, D J (1961) Territorial behaviour in barnacle settlement. Journal of Experimental 
Biology 38(2): 429. 
Crisp, D J; Austin, A P (1960) The acti on of copper in an ti-fouling paints. Annals of Applied 
Biology 48: 787-799. 
Crisp, D J; Meadows, P S (1962) The chemical  basis of gregariousness in cirripedes. 
Proceedings of the Royal Society of London Series B . 158: 364. 
Crisp, D J; Walker, G; Yule, A B (1985) Adhesion and substrate choice in mussels and 
barnacles. Journal of Colloid and Interface Science  104(1): 40-50. 
Floerl, O; Inglis, G J (2003) Boat ha rbour design can exace rbate hull fouling. Austral 
Ecology. 28: 116-127. 
Floerl, O; Inglis, G J; Marsh, H M (2005) Selectivity in vector  management: an investigation 
of the effectiveness of measures used to pr event transport of Non- indigenous species. 
Biological Invasions  7: 459-475. 
Foster, B A (1971) Desiccation as a factor in the intertidal zonation of barnacles. Marine 
Biology 8: 12-29. 
Geller, J; Sotka, E E; Kado, R; Palumbi, S R;  Schwindt, E (2008) Sour ces of invasion of a 
northeastern Pacific acorn barnacle, Balanus glandula , in Japan and Argentina. Marine 
Ecology Progress Series  358: 211-218. 
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  47  
Gren, I-M; Isaacs, L; Carlsson, M (2007) Calculations of Costs of Invasive Alien Species in 
Sweden—Technical Report. Working paper no. 7,.  Department of Economics, Swedish 
University of Agricultural Sciences. Uppsala, Sweden.  
Gulf States Marine Fi sheries Commission (2005) Balanus trigonus  fact sheet. 
http://nis.gsmfc.org/nis_factsheet 2.php?toc_id=121 - accessed 21 May 2008. 
Hung, O S; Thiyagarajan, V; Qia n, P Y (2008) Preferential attach ment of barnacle larvae to 
natural multi-species biofilms: Does surface wetability matter? Journal of Experimental 
Marine Biology and Ecology 361: 36-41. 
ICES (2004) Report of the Working Group on Introductions and Transfers of Marine 
Organisms (WGITMO). 25-26 March 2004; Cesenatico, Italy. 
Inglis, G J; Floerl, O; Unwi n, M; Ponder-Sutton, A; Seawar d, K; Kospartov, M; Bell, A; 
Kluza, D (in prep.). The biosecurity risks associated with biofouling on international 
vessels arriving in New Zealand. Summary of the patterns and predictors of fouling.  
Biosecurity New Zealand Technical Paper. 
Jarrett, J N (1997) Temporal variatio n in substratum specificity of Semibalanus balanoides  
(Linnaeus) cyprids . Journal of Experimental  Marine Biology and Ecology  211: 103-
114. 
Jeffery, C J (2002) New settlers and recruits do not enhance settlement of a gregarious 
intertidal barnacle in New South Wales. Journal of Experiment al Marine Biology and 
Ecology 275:131-146. 
Kado, R (2003) Invasion of Japanese shores by the NE Pacific barnacle Balanus glandula  and 
its ecological and biogeographical impact. Marine Ecology Progress Series  249: 199-
206. 
Kent, A; Hawkins, S J; Patrick Doncaster, C (2003) Population consequences of mutual 
attraction between settling and adult barnacles. Journal of Animal Ecology  72: 941-952. 
Kospartov K; Inglis G; Seawar d K; Brink A van den; D’Arch ino R; Ahyong S (in press) Non-
indigenous and cryptogenic mari ne species in New Zealand . Current state of knowledge 
– Interim report prepared for MAFBNZ Post Border Directorate. MAF Biosecurity New 
Zealand. Wellington, New Zealand. 
Laird, M C; Griffiths, C L (2008) Present di stribution and abundan ce of the introduced 
barnacle Balanus glandula  in South Africa. African Journal of Marine Science 30(1): 
93-100. 
Lawson, J; Davenport, J; Whitaker, A (2004)  Barnacle distribution in Lough Hyne Marine 
Nature Reserve: a new baseline and an  account of invasion by the introduced 
Australasian species Elminius modestus  Darwin . Estuarine, Coastal and Shelf Science 
60 (4): 729-735. 
Leppakoski, E; Olenin, S (2000) Non-native species and rates of spread: lessons from the 
brackish Baltic Sea. Biological Invasions  2: 151-163. 
Lewis, J (2004) Hull Fouling as a Vector for the Trans location of Marine Organisms. Report 
1: Hull fouling research.  Ballast W ater Research Seri es Report No. 14. Dept of 
Agriculture, Fisheries and Fo restry; Canberra, Australia; 129pp. 
Lohrer, A M; Whitlatch, R B; Wada, K; Fukui , Y (2000) Home and Away: comparisons of 
resource utilization by a marine speci es in native and invaded habitats. Biological 
Invasions  2: 41-57.  
Lopez, D A; Lopez, B A; Burgos, I C; Arriagada, S E; Gonzalez, M L (2007) Consequences 
of base modification in hummocks of the barnacle Austromegabalanus psittacus . New 
Zealand Journal of Marine and Freshwater Research  41: 291-298. 
Miller, L (2007) Feeding in extreme flows: beha viour compensates for mechanical constraints 
in barnacle cirri. Marine Ecology Progress Series  349: 227-234. 
Minchin, D; Gollasch, S (2003) Fouling and sh ips’ hulls: how changing circumstances and 
spawning events may result in th e spread of exotic species. Biofouling  19 (supplement): 
111-122. 
48  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
NIMPIS (2010). Megabalanus rosa reproduction a nd habitat. National Introduced Marine Pest 
Information System. Web publication http://www.marinepests.gov.au/nimpis . Date of 
access: 30/08/2009 
O’Riordan, R M; Ramsay, N F (1999) The cu rrent distribution a nd abundance of the 
Australasian barnacle Elminius modestus in Portugal. Journal of the Marine Biological 
Association of the United Kingdom  79: 937-939. 
Otani, M; Oumi, T; Uwai, S; Hanyuda, T; Pr abowo, R E; Yamaguchi, T; Kawai, H (2007) 
Occurrence and diversity of barnacles on international sh ips visiting Osaka Bay, Japan, 
and the risk of their introduction. Biofouling  23: 277-286.  
Perez-Losada, M; Hoeg, J T; Cr andall, K A (2004) Unravelling the Evolutionary Radiation of 
the Thoracican Barnacles Using Molecula r and Morphological Evidence: A Comparison 
of the Several Divergence Time Estimation Approaches. Systematic Biology  53(2): 244-
264. 
Piola, R F; Johnston, E L (2008) The potential fo r translocation of mari ne species via small-
scale disruptions to anti-fouling surfaces. Biofouling  24(3): 145-155. 
Railkin, A I (2004) Marine Biofouling: Colonization Processes and Defences . CRC Press; 
303pp 
Railkin, A I; Makarov, V N; Shoshina, E V (1985) The influence of s ubstrate on settlement 
and attachment of Laminaria saccharina  zoospores. Biologija Morya  1: 37 
Ruppert, E E; Fox, R S; Barnes, R D (2004) Invertebrate Zoology: A functional evolutionary 
approach (7th edition). Thompson Lear ning; Victoria, Australia. 
Sakaguchi, I; Namihira, T (2006) Settlement and growth of barnacles at  a power station in 
Okinawa. Sessile Organisms  23(1): 17-23.  
Schwindt, E (2007) The invasion of the acorn barnacle Balanus glandula  in the south-western 
Atlantic 40 years later. Journal of the Marine Biologi cal Association of the United 
Kingdom  87:1219-1225. 
Simon-Blecher, N; Granevitze, Z; Achituv, Y (2008) Balanus glandula: from North-West 
America to the west coast of South Africa. African Journal of Marine Science  30(1): 
85-92. 
Skerman, TM (1960) Ship-fouling in New Zeal and waters: a survey of marine fouling 
organisms of the coastal and overseas trades. New Zealand Journal of Science  3: 620-
648. 
South Carolina Department of Natural Resources (2006) News release #06–342: Pesky exotic 
barnacle found in Charleston Ha rbor estuary. Accessed 23 May 2008. 
Species 2000. Catalogue of Life: Dynamic Checklist . 
http://www.catalogueoflife.org/dynamic-ch ecklist/browse_taxa.php. Accessed 3 March 
2009. 
Thiyagarajan, V; Venugopalan, V P; Subramoniam, T; Nair, V K (1997) Macro-fouling in the 
cooling water conduits of a coastal power station. Indian Journal of Marine Sciences  
26(3): 305-308. 
Thiyagarajan, V; Harder, T; Qian, P Y (2002a ) Effect of the physiological condition of 
cyprids and laboratory-mimicked seasonal conditions on the metamorphic success of 
Balanus amphitrite  Darwin. Journal of experimental Marine Biology and Ecology  274: 
65-74. 
Thiyagarajan, V; Harder, T; Qian, P Y (2002b) Relationship between cyprid energy reserves 
and m
etamorphosis in the barnacle Balanus amphitrite  Darwin (Cirripedia; Thoracica). 
Journal of experi mental Marine Biology and Ecology  280: 79-93. 
Thiyagarajan, V; Harder, T; Qian, P Y (2003) Co mbined effect of temperature and salinity on 
larval development and attachme nt of the subtidal barnacle Balanus trigonus  Darwin. 
Journal of experi mental Marine Biology and Ecology  287: 223-236. 
Watson, D I; O'Riordan, R M; Barnes, D K A;  Cross, T (2005) Temporal and spatial 
variability in the recruitment of barnacles and the local dominance of Elminius 
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  49  
50  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis modestus  Darwin in SW Ireland.  Estuarine, Coastal and Shelf Science  63(1-2): 119-
131. 
Wood, E J F; Allen, F E (1958) Common marine fouling organisms of  Australian waters. 
Department of the Navy, Navy Office; Melbourne, Australia; 23 pp. 
Wiegem ann, M (2008) Wild cyprids metamorp hosing in vitro reveal the presence of Balanus 
amphitrite Darwin, 1854 in the German Bight basin. Aquatic Invasions  3(2): 235-238. 
Yamaguchi, T; Prabowo, R E; Ohshiro, Y; Shim ono, T; Jones, D; Kawai, H; Otani, M; 
Oshino, A; Inagawa, S; Akaya, T; Tamura , I (2009) The introduction to Japan of the 
Titan barnacle, Megabalanus coccopoma (Darwin, 1854) (Cirripedia: Balanomorpha) 
and the role of shipping in its translocation. Biofouling  25: 325-333.  
Zabin, C J; Altieri, A (2007) A Hawai’ian limpet  facilitates recruitmen t of a competitively 
dominant barnacle. Marine Ecology Progress Series  337: 175-185. 
Zabin, C J; Zardus, J; Pitombo, F B; Fread, V; Hadfield, M G (2007) A tale of three seas: 
consistency of natural history traits in a Caribbean-Atlantic barnacle introduced to 
Hawai’i. Biological Invasions  9: 523-544. 
Zardus, J D; Hadfield, M G (2005) Multiple origin s and incursions of the Atlantic barnacle 
Chthamalus proteus  in the Pacific. Molecular Ecology  14: 3719-3733. 
Zvyagintsev, A Y; Korn, O M (2003) Life history of the barnacle Balanus amphitrite  Darwin 
and its role in fouling communities of Peter the Great Bay, Sea of Japan. Russian 
Journal of Marine Biology  29(1): 41-48.  
4.4. BIVALVE MOLLUSCS  (CLASS: BIVALVIA) 
4.4.1. Hazard Identification 
The class Bivalvia are m
olluscs characteris ed by the presence of two calcareous 
valves which enclose soft body parts. Bivalves are found in aquatic environments, 
ranging from freshwater to marine, from the in tertidal to abyssal depths.  In general, 
bivalves live in benthic environments with ei ther an epifaunal or infaunal habit.  The 
majority of bivalve species have limited m obility and therefore are largely sedentary 
and suspension feeding.  Bivalves are ofte n important ecosystem engineers providing, 
for example, benthic pelagic coupling and habi tat structure. This in turn influences 
local community diversity, abundance and car rying capacity in both the benthic and 
pelagic environments (Creese et al. 1997, Newell 2004, Porter et al. 2004, Ruesink et 
al. 2004, Chapman et al. 2005, Norling & Kautsky 2007). For example, a clump of mussels may accumulate sediments in their byssal threads which provide habitat for species such as polychaetes.  These musse ls may also provide nutrients through the 
deposition of faecal and pseudofaecal pellets which will sustain associated organisms 
(Kochmann et al. 2008).  Many bivalve species are of high economic a nd social importance as food species.  
Naturally occurring populations, enhanced popul ations and aquacultu re all contribute 
to produce bivalves for human consumpti on.  Bivalve species which are commonly 
cultured demonstrate characte ristics of rapid growth, hi gh fecundity and dominant 
occupation of space (e.g. Mytilus  spp. Bayne 1976). These characteristics are 
particularly common to oysters and mussels  that act as ecosystem engineers.  
However, these characteristics also give many bivalve species the potential to be 
highly invasive.    Known invasive marine species include the mussels Perna perna (Hicks & Tunnell 
1995), Perna viridis  (Rajagopal et al. 2006), Mytilus galloprovincialis  (Hanekom 
2008), Mytilopsis sallei  (Galil & Bogi 2008) and the oyster Crassostrea gigas .  
Invasions by these species have been reported to have significant impacts on 
ecosystem structure and function (Bow nes & McQuaid 2006, Ruesink et al. 2006, 
Kochmann et al. 2008).  Invasi ve bivalves have also had significant economic impacts 
through modifying fisheries (e.g. mussel cultu re in South Africa (Robinson et al. 
2005), oyster culture in New Zealand (Dinaman i 1991) and fouling infrastructure (e.g. 
Perna viridis in India (Rajagopal et al. 1991)).  Bivalves are well known components of vessel biofouling (Skerman 1960, Lewis 
2004, Farrapeira et al. 2007) .  Bivalves, notably Mytilid  species, have been frequently 
reported to be the dominant fouling macr o-organisms on vessel hulls (Apte et al. 
2000, Gollasch 2002, Farrapeira et al. 2007, Lee & Chown 2007).  Recreational 
vessels have been implicated in at least tw o significant bivalve incursions in Australia 
with P. viridis in tropical far north Queensla nd (Stafford et al. 2007) and the 
establishment of M. sallei  in Darwin harbour (Willan et  al. 2000). Differences exist 
between vessel type and the likely bivalv e fouling assemblages, with recreational 
vessels more likely to be fouled with oys ters and commercial vessels with mussels 
(James & Hayden 2000, Inglis et al. in prep.).   
 
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  51  
 
Most temperate bivalve species known to be pests (e.g. C. gigas , M. galloprovincialis  
and Musculista senhousia ) have previously been found in New Zealand (Dinamani 
1971, Creese et al. 1997, Hayward et  al. 1999b, Gardner 2004). Although P. viridis 
and M. sallei  have established in various global locations, including eastern Australia, 
they are not yet established in New Zealand.  P. perna,  which is native to South 
Africa and a known pest in central and southe rn America, has been  introduced to New 
Zealand but not yet known to ha ve established (MAFBNZ 2009).  Dreissena 
polymorpha is also a high profile, biofouling, pest bivalve, however, it is restricted to 
freshwater environments and as a consequence is unlikely to reach New Zealand as vessel biofouling.    Given that bivalves: 
 are known components of marine  biofouling assemblages;  
 are known to have been introduced to new locations; 
 have known impacts on core values; 
it is concluded that bivalve species represent a potential hazard via vessel biofouling. 
 
4.4.2. Risk Assessment 
4.4.2.1.  Entry assessment 
Typical fouling representatives of Biva lvia are of
ten dominant components of 
intertidal and shallow subtidal coastal ecosystems.  Most bivalves produce large 
quantities of gametes, which have a plankt onic larval stage enabling dispersal beyond 
the parent location. Therefore, propagules of bivalve species with th e potential to foul 
artificial structures are li kely to be present in the majority coastal waters.   
 
Anti-fouling systems can be effective in kil ling, or deterring the settlement of, bivalve 
larvae.  However, anti-fouling systems have been implicated in selecting specifically 
for non-indigenous species (Dafforn et al. 2008) which have resist ance to the active 
ingredients.  For example, bivalve species found in polluted environments, such as 
ports, are likely to have de veloped a level of resistance to copper compounds (Hoare 
et al. 1995 a, b).  As a result, vessels exposed to competent bivalve larvae in 
port/marina environments could become fouled regardless of the quality of their anti-
fouling system.    Source bivalve populations may thrive due to their ability to survive extended periods 
in conditions unsuitable for potential comp eting species. For example, many marinas 
are situated in estuarine locations which can often harbour wide environmental ranges.  Apte et al. (2000)  described the spawning of M. galloprovincialis  on the hull 
of a naval vessel following its arrival to Hawai’i despite the vessel having spent 9 
days moored in the Columbia river as freshw ater treatment (salinity 2-10 ppt) prior to 
its departure.    Whether attached by byssus or cemented dir ectly to the surface, once adhered to a 
surface bivalves can resist substantial dislodgement forces.  However, many vessel 
owners/operators are likely to remove  obvious macro-fouling to minimise 
performance penalties.  As bivalve fou ling is likely to be conspicuous on, and 
consequently removed from, the open hull, bi valves attached to vessels en-route are 
52  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
likely to be in locations not easily inspected and/or cleaned such as the propeller, 
drive shaft or sacrificial anodes.  Biva lves are well known to survive travel over 
considerable distances with th e widespread distribution of M. galloprovincialis  
attributed primarily to transport as biofou ling.  Internal seawater systems, including 
seachests, are also likely locations for bi valve fouling (Lewis & Coutts 2009).   
 Given that particular bivalve species: 
 are known to occur along vessel pathways; 
 are known to be associated with vessel biofouling; 
 are known to survive translocation; 
it is concluded that the likelihood of entry of non-indigenous bivalv e species into New 
Zealand via vessel biofou ling is non-negligible. 
 
4.4.2.2.  Establishment assessment 
As bivalves are likely to be  stim
ulated to spawn due to environmental stressors such 
as thermal shock (e.g. Apte et al. 2000), the hi ghest risk areas are likely to be the first 
ports of call vessels make following arrival in the coastal zone.  The ports of first 
arrival in New Zealand provide a range of natural and artifi cial habitats that include 
hard to soft and natural to  artificial substrates.  Te mperature ranges are generally 
warm temperate in the north and cool temp erate in the south, although most ports are 
exposed to significant fluctuations in envir onmental conditions due  to their estuarine 
location.  These conditions are likely to be conducive to successful spawning, larval 
development and settlement for bivalve species  arriving from the majority of potential 
donor ports.  However, only the northern most ports (north of Auckland) are likely to 
provide conditions where tr opical species, such as P. viridis, could establish persistent 
populations.   
 Areas to the north of the North Island and at the top of the South Island provide 
conditions which are conducive to bivalve aquaculture.  Neighbouring major shipping 
routes, these areas are also conducive to  the establishment of non-indigenous 
bivalves.  Whilst farmers may identify and remove novel species, the cryptic 
morphology of invasive bivalves compar ed to native species may mean a new 
incursion is undetected by farmers.  Past in cursions have also s hown that aquaculture 
areas can perpetuate non-targe t invasive species which can then be vectored by stock 
and equipment movements, for example, Didemnum vexillum, Styela clava and 
Eudistoma elongatum . 
 On arrival to New Zealand, non-indigenous bivalve species could establish a population by relocation from the vessel to the receiving environment.  Most adult 
bivalves are unlikely to actively relocate from a vessel hull, however, they are known 
to survive physical removal (MAFBNZ 2009). This is especially evident in species 
that attach by byssus.  Species without by ssus, such as Pacific oysters, are also known 
to survive physical removal and live in ha bitat that would otherwise be unsuitable. 
However, such species are vulnerable to damage of the attachment surface which exposes the soft body to damage, disease and predation.    Bivalves may also establish in the receivi ng environment by gamete release (Apte et 
al. 2000).  Bivalves produce large numbers of gametes (in some cases millions of eggs per female), usually released in a synchronous broadcast spawn.  For broadcast 
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  53  
spawning species to successf ully reproduce they would need the following conditions 
to be met: the presence of at least one bivalve of each se x, the bivalves to be in a 
reproductively mature condition, to be close enough together for spawned eggs and 
sperm to fertilise, and to be stimulated to spawn synchronously (e.g. oysters are known to mutually stimulat e spawning (Pauley 1988)). 
 The likelihood of the above conditions being met is high, as bivalves are generally gregarious settlers forming clumps rather than existing as isolated individuals. The 
existence of bivalves in clumps will enable mature adults to detect chemosensory spawning signals leading to greater like lihood of spawning success. Although anti-
fouling compounds can reduce the condition of adult oysters, such as C. gigas, the 
viability and growth of embryos or larvae is unaffected (His & Robert 1987).  In 
addition, physical damage, such as that su stained during removal, may result in the 
release of gametes.  Further, New Zeal and ports are generally located in highly 
productive estuarine waters.  Bivalve larvae are usually free swimming with the planktonic stage lasting from days 
to months.  Thus larvae may move well beyond the environment in which they were released.  In the case of the port or mari na environment, this may considerably 
enhance the chances of larvae finding a su itable habitat for colonisation.  While 
mortality of the larvae can be high (>90%) th e vast quantities of larvae released mean 
that a single breeding pair can still establish a founder population. 
 Some bivalve species brood larvae by inhaling sperm and hold the fertilised eggs until they are released as the swimming veliger st age.  Therefore, a single brooding female, 
fertilised in a previous port, could effectively establish a new population.   
 Given that particular bivalve species: 
 are known to inhabit similar climates/env ironments to those that exist in New 
Zealand; 
 are capable of forming establishi ng populations in New Zealand; 
 have been recently recorded as  established in New Zealand; 
it is concluded that the likelihood of esta blishment of non-indige nous bivalve species 
in New Zealand via vessel biofouling is non-negligible.  
4.4.2.3.  Consequence assessment 
An incursion of a non-indigenous bivalv e could have significant im
pacts on 
economic, environmental, social and cultural values.  Key to this is the ecosystem 
engineering nature of many bivalve species, i.e. the modification of habitat for their 
own benefit (Ruesink et al. 2006).  Reported ecosystem impacts of bivalves include 
the alteration of water quality (Porter et al. 2004), displacement of  native species (e.g. 
M. galloprovincialis  in South Africa) (Hanekom  2008), the modification of 
community structure (Chapman et al. 2005) and the modification of ecosystem 
function (Norling & Kautsky 2007).  Bivalves can exert both top-down and bottom-up 
influences over ecosystem function by filtering phytoplankton and providing (or removing) nutrients from the water column or substrate.  However, an overabundance 
of bivalves can lead to deoxygena tion of sediments, inhibition of 
nitrification/denitrification processes, release of phosphorus into the water column 
and build-up of hydrogen sulphide  in sediments (Newell 2004). 
54  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
 
The Asian clam, Corbula amurensis, was introduced into th e northern San Francisco 
Bay in 1986. Within two years the clam predominated (>95%) in both total number of individuals (>25,000 individuals per m
2) and biomass and had halted the re-
establishment of dry-period es tuarine species (Nicholls et  al. 1990). As a result, the 
benthic community structure and dynamics of  this region of San Francisco Bay have 
been irreversibly altered. For example, filtering of the water column by  the large clam 
populations has eliminated summe r-long phytoplankton blooms since 1987 
(Kimmerer 2006). Such changes to primar y productivity has resulted in varying 
responses, for example, mysid abundance de clined sharply, and populations of three 
species of estuarine-dependent  fish declined, however, other species have maintained 
their abundance (Kimmerer 2002 In Kimmerer 2006). In addition, the altered food 
chains have halted the commercial harvesting of anchovies in local waters (Kimmerer 
2006).   In Australia, the non-indigenous clam Corbula gibba to Port Phillip Bay (Australia) 
has reached densities of up to 2600 per m
2 and dominates deeper subtidal assemblages 
(Currie & Parry 1996 In Talman and Keough 2001). It has been shown in field 
experiments that ambient densities of C. gibba can significantly affect the size and 
growth of juveniles of the commercial scallop Pecten fumatus via resource 
competition (Talman and Keough 2001).    In the USA, the green mussel ( P. viridis) was first identified in Tampa Bay, Florida, 
in 1999. Since then dense assemblages of thes e bivalves now clog sea water intakes of 
power stations and desalination plants , weigh down navigational buoys and foul 
vessel hulls and engine s (Baker et al. 2004 In Mitchem et al. 2007). 
 Recreational fisheries are al so vulnerable to non-indigenous bivalve species.  For 
example, the introduction of the clam Ruditapes philippinarum  to Venice, Italy, has 
been linked to reduced fishing catch rates,  and fishers in turn have changed their 
angling habits (Albernini et al. 2007).  The impacts of introduced bivalves in Ne w Zealand have generally been poorly 
described.  For example, the Pacific oyster ( C. gigas) displaced the native oyster 
(C. glomerata) from a broad distribution in the inte rtidal to a rare occurrence in the 
high intertidal. However, the implications of this shift on ecosystem structure and 
function are unknown.  In economic terms, the Pacific oyster has become the only commercially cultured oyster in New Zeal and because its rapid growth and high 
fecundity lead it to out-compe te native oysters on growing racks.  While this has 
provided New Zealand oyster farmers with an  easily marketable species it has also 
meant they have lost the opportunity to cu ltivate and market a niche product as has 
been successfully demonstrated with the New Zealand green lipped mussel ( Perna 
canaliculus). Equally northern M āori have been deprived of a toanga species. 
 The introduction of M. senhousia  to New Zealand has resulted in significant 
alterations in benthic community struct ure, composition and abundance associated 
with the ephemeral M. senhousia  mats (Creese et al. 1997, Hayward et al. 1999a).  In 
the Waitemata harbour, not only can M. senhousia  produce substantial mats, but they 
also remain in lower densities throughout the harbour, contributi ng to changes in the 
benthic species composition (Lohrer  et al. 2008).  The impact of M. senhousia  could 
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  55  
be both beneficial, by providing an abundant food source to higher trophic levels, and 
detrimental, by reducing the diversity of native species upon which ecosystem 
resilience may have been reliant. 
 New Zealand has a substantial aquaculture industry heavily reliant  on the culture of 
the native mussel P. canaliculus .  This could be at risk with the introduction of a 
competing bivalve species.  Given the reliance upon a single species, the New 
Zealand industry could be vulnerable if a successfully introduced mussel caused a reduction in productivity similar to th at already caused by the blue mussel ( M. 
galloprovincialis ).  The success of P. canaliculus as an aquaculture species can be 
linked to qualities which make it a potentia l invasive in other countries (Pollard & 
Hutchings 1990).  These features could provide P. canaliculus  with a substantial 
capacity to resist invasive mussels, how ever, this may be limited to certain 
environments as noted for P. perna resisting M. galloprovincialis  in South Africa 
(Nicastro et al. 2008).  In addition to substantial econom ic effects, any loss or 
reduction in abundance of P. canaliculus will also have impacts on cultural and social 
values as it is a significant  target species for recreational and customary harvesters. 
 Bivalves may facilitate the import of associ ated nuisance flora and fauna that are not 
established in New Zealand.  For example,  fouling species on mussel shells may 
reduce the profit margin for in-shell pr oducts due to the need for additional 
processing.  Other associated species could foul aquaculture  infrastructure or wild 
fishery habitats impacting of productivity, accessi bility and/or desirability of harvests. 
4.4.3. Risk estimation 
Given that particular bivalve species: 
 are comm 
only associated with hull fouling; 
 have been introduced and established globall y via vessel fouling; 
 have known invasive behaviour; 
 can be ecosystem engineers; 
 can facilitate the entry a nd establishment of a variety of other non-indigenous 
organisms; 
 have been found to cause substantive change to native ecosystems globally, 
including New Zealand; 
 may negatively impact green lipped mu ssel aquaculture which accounts for the 
majority of New Zealand’s aquaculture earnings, in addition to customary and 
recreational harvesting; 
 that are associated with high impacts, such as Perna perna, Perna viridis, 
Mytilopsis sallei , are yet to establish in New Zealand; 
it is concluded that non-indigenous bivalv es pose a non-negligib le risk to New 
Zealand’s coastal waters. 
 
4.4.4. Risk management 
The trea
tment of hulls to manage the risks posed by bivalves is covered by the generic 
options (Section 3).  Specific considerations for hull fouling bivalves should include 
appropriate local containment.  While it ma y be possible to remove large specimens 
or clumps by hand, caution should be exerci sed due to the likelihood of release of 
56  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
associated non-indigenous organisms in to the receiving environment. Moreover, 
physical damage may cause the accidental re lease of gametes which could fertilise 
and subsequently found a population.   
 
4.4.5. References 
Alberini, A; Zanatta, V; Rosato, P (2007) Com
bining actual and contingent 
behaviour to estimate the value of spor ts fishing in the Lagoon of Venice. 
Ecological Economics , 61: 530-541. 
Apte, S; Holland, B S; Godwin, L S; Gardne r, J P A (2000) Jumping ship: a stepping 
stone event mediating tran sfer of a non-indigenous species via a potentially 
unsuitable environment.  Biological Invasions  2: 75-79. 
Bayne, B L (1976) Marine mussels: their ecology and physiology . Cambridge 
University Press; Cambridge, United Kingdom. 
Bownes, S J; McQuaid, C D (2006) Will the invasive mussel Mytilus 
galloprovincialis Lamarck replace the indigenous Perna perna  L. on the south 
coast of South Africa?  Journal of Experimental Marine Biology and Ecology 
338: 140-151. 
Chapman, M G; People, J; Blockley, D (2005) Intertidal assemblages associated with 
natural corallina  turf and invasive mussel beds. Biodiversity and Conservation  
14: 1761-1776. 
Creese, R; Hooker, S; De Luca, S; Wh arton, Y (1997) Ecology and environmental 
impact of Musculista senhousia  (Mollusca: Bivalvia: Mytilidae) in Tamaki 
Estuary, Auckland, New Zealand. New Zealand Journal of Marine and 
Freshwater Research  31: 225-236. 
Crow, S P (1979) Marine fouling on the Ma ui A platform.  Unpublished thesis, The 
University of Auckland, Auckland. 
Dafforn, K A; Glasby, T M; Johnston, E L ( 2008) Differential effects of tributyltin 
and copper antifoulants of recrui tment of non-indigenous species. Biofouling  
24: 23-33. 
Dinamani, P (1971) Occurrence of the Japanese oyster, Crassostrea gigas  
(Thunberg), in northland, New Zealand.  New Zealand Journal of Marine and 
Freshwater Research  5: 352-357. 
Dinamani, P (1991) Introduced Pacific oysters  in New Zealand. In: Leffler, M; Greer, 
J R (Eds) The ecology of Crassostrea gigas  in Australia, New Zealand, France 
and Washington . Maryland Sea Grant, College  Park, Maryland. VM-SG-TS-
92-07. 
Farrapeira, C M R; de Melo, A V de O M; Barbosa, D F; da Silva, K M E (2007) Ship 
hull fouling in the port of Recife, Pernambuco. Brazilian Journal of 
Oceanography  55: 207-221. 
Gardner, J P A (2004) A historic al perspective of the genus Mytilus (Bivalvia: 
Mollusca) in New Zealand: multivariate morphometric analyses of fossil, midden and contemporary blue mussels. Biological Journal of the Linnean 
Society  82: 329-344. 
Galil, B S; Bogi, C (2008) Mytilopsis sallei (Mollusca: Bivalvia: Dreissenidae)  
established on the coast of Israel. JMBA 2 Biodiversity Records  Published 
Online 
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  57  
Glasby, T M (1999) Differences between subtidal epibiota on pier pilings and rocky 
reefs at marinas in Sydney, Australia. Estuarine, Coastal and Shelf Science 48: 
281-290. 
Gollasch, S (2002) The importance of ship hull fouling as a v ector of species 
introductions into the North Sea. Biofouling  18: 105-121. 
Hanekom, N (2008) Invasion of an indigenous Perna perna mussel bed on the south 
coast of South Africa by an alien mussel Mytilus galloprovincialis  and its 
effect on the associated fauna. Biological Invasions  10: 233-244. 
Hayward, B W; Morley, M S; Hayward, J J;  Stephenson, A B; Bl om, W M; Hayward, 
K A; Grenfell, H R (1999a) Monitoring studies of the benthic ecology of 
Waitemata Harbour, New Zealand. Records of the Auckland Museum  36: 95-
117. 
Hayward, B W; Morley, M S; Stephenson, A B; Blom, W M; Grenfell, H R; Prasad, 
R; Rogan, D; Thompson, F; Cheetham, J; Webb, M (1999b) Intertidal and 
subtidal biota and habitats of the central Waitemata Harbour .  Technical 
Publication 127, Auckland Regiona l Council; Auckland, New Zealand. 
Hicks, D W; Tunnell, J W (1995) Ecological notes and patterns of  dispersal in the 
recently introduced mussel, Perna perna (Linne, 1758), in the Gulf of Mexico. 
American Malacological Bulletin 11: 203-206 
His, E; Robert, R (1987) Comparative effect s of two anti-fouling paints on the oyster 
Crassostrea gigas .  Marine Biology 95: 83-86. 
Hoare, K; Beaumont, A R; Davenport, J (1995a) Variation am ong populations in the 
resistance of Mytilus edulis  embryos to copper: adaptation to pollution? 
Marine Ecology Progress Series 120: 155-161. 
Hoare, K; Davenport, J; B eaumont, A R (1995b) Effects of exposure and previous 
exposure to copper on growth of ve liger larvae and survivorship of Mytilus 
edulis  juveniles. Marine Ecology Progress Series 120: 163-168. 
Inglis, G J; Floerl, O; Unwi n, M; Ponder-Sutton, A; Seaward, K; Kospartov, M; Bell, 
A; Kluza, D (in prep.). The biosecurity risks associ ated with biofouling on 
international vessels arriving in New Zealand. Summary of the patterns and 
predictors of fouling.  Biosecurity New Zealand Technical Paper. 
James, P; Hayden, B (2000) The potential for the introduction of exotic species by 
vessel hull fouling: a preliminary study . NIWA client report WLG00/51. 
Kimmerer, W J (2006) Response of anchovies dampens effects of the invasive bivalve 
Corbula amurensis on the San Francisc o Estuary foodweb. Marine Ecology 
Progress Series , 324: 207-218. 
Kochmann, J; Buschbaum, C; Volkenborn, N;  Reise, K (2008) Shift from native 
mussels to alien oysters: Differentia l effects of ecosystem engineers. Journal 
of Experimental Mari ne Biology and Ecology  364: 1-10. 
Lee, J E; Chown, S L (2007) Mytilu s on the move: transport of an invasive bivalve to 
the Antarctic.  Marine Ecology Progress Series  339: 307-310. 
Lewis, J A (2004) Hull fouling as a vector fo r the translocation of marine organisms: 
Report 1: Hull fouling research.  Ballast Water Research Series Report No. 14 
Department of Agriculture, Fisheries and Forestry; Canberra, Australia. 
Lewis, J A; Coutts, A D M (2009) Chapter 24 Biofouling Invasions. In: Dürr, S; 
Thomason, J (Eds) Biofouling Wiley-Blackwell. 
Lohrer, D; Townsend, M; Morrison, M; Hewitt, J (2008) Change in the benthic 
assemblages of the Waitemata Harbour : Invasion risk as a function of 
community structure.   Biosecurity New Zealand Technical Paper No: 2008/17 
MAF Biosecurity New Zealand; Wellington, New Zealand. 
58  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
MAFBNZ (2009) Perna perna  http://www.biosecurity.govt .nz/pests/perna-perna  
Accessed 20 April 2009. 
Mitchem, E L: Fajans, J S; Baker, S M (2007). Contrasting response of two native 
crustaceans to non-indigenous prey, the green mussel, Perna viridis . Florida 
Scientist  70(2), 180-188. 
Nicastro, K R; Zardi, G I; McQuaid, C D (2008) Movement  behaviour and mortality 
in invasive and indigenous mussels: re silience and resistance strategies at 
different spatial scales.  Marine Ecology Progress Series  372: 119-126. 
Newell, R I E (2004) Ecosystem influences of natural and cultiv ated populations of 
suspension-feeding bivalve molluscs: a review. Journal of Shellfish Research  
23: 51-61. 
Nichols FH, Thompson JK and Schemel LE (1990) Remarkable invasion of San 
Francisco Bay (California, USA) by the Asian clam Potamocorbula amurensis. 
II. Displacement of a former commun ity. Marine Ecology Progress Series 66: 
95–101 
Norling, P; Kautsky, N (2007) Structural and functional effects of Mytilus edulis  on 
diversity of associated speci es and ecosystem functioning.  Marine Ecology 
Progress Series  351: 163-175. 
Pauley, G B; Van Der Raay, B; Troutt, D (1988) Species profiles: life histories and 
environmental requirements of coastal fishes and invertebrates (Pacific 
Northwest)--Pacific oyster .  U.S. Fisheries and Wildlife Service Biological 
Report 82(11.85). U.S. Army Corps of Engineers, TR EL-82.4. 
Pimentel, D; Zuniga, R; Morrison, D ( 2005) Update on the environmental and 
economic costs associated with alien-i nvasive species in the United States. 
Ecological Economics  52: 273-288. 
Pollard, D A; Hutchings, P A (1990) A review  of exotic marine organisms introduced 
to the Australian region. II.  Invertebrates and algae. Asian Fisheries Science  
3:223-250. 
Porter, E T; Cornwell, J C; Sanf ord, L P (2004) Effect of oysters Crassostrea 
virginica  and bottom shear velocity on bent hic-pelagic coupling and estuarine 
water quality.  Marine Ecology Progress Series  271: 61-75. 
Rajagopal, S; Sasikumar, N; Azariah, J;  Nair, K V K (1991) Some observations on 
biofouling in the cooling water con duits of a coastal power plant. Biofouling  3: 
311-324. 
Rajagopal, S; Venugopalan, V P; van der Veld e, G; Jenner, H A (2006) Greening of 
the coasts: a review of the Perna viridis success story. Aquatic Ecology 40: 
273-297. 
Robinson, T B; Griffiths, C L; McQuaid, C D;  Rius, M (2005) Marine alien species of 
South Africa - status and impacts. African Journal of Marine Science  27: 297-
306. 
Ruesink, J L; Feist, B E; Harvey, C J; H ong, J S; Trimble, A C;  Wisehart, L M (2006) 
Changes in productivity associated with  four introduced species: ecosystem 
transformation of a ‘pristine’ estuary.  Marine Ecology Progress Series  311: 
203-215. 
Skerman, T M (1960) Ship fouling in New Zealand waters: A survey of marine 
fouling organism
s from vessels of  the coastal and overseas trade. New Zealand 
Journal of Science  3: 620-648. 
Stafford, H; Willan, R C; Neil, K M ( 2007) The invasive Asian Green Mussel, Perna 
viridis  (Linnaeus, 1758) (Bivalvia: Mytilidae), breeds in Trinity Inlet, tropical 
northern Australia.  Molluscan Research  27: 105-109. 
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  59  
60  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis Willan, R C; Russel, B C; Murfet, N B; M oore, K L; McEnnulty, F R; Horner, S K; 
Hewitt, C L; Dally, G M; Campbell, M L; Bourke, S T (2000) Outbreak of 
Mytilopsis sallei (Récluz, 1849) (Bival via: Dreissenidae) in Australia. 
Molluscan Research 20: 25–30. 
   
4.5. BRISTLEWORMS  (CLASS: POLYCHAETA) 
4.5.1. Hazard Identification 
Bristleworms (polychaetes) are the largest class of wor
ms (a nnelids) with more than 13,000 
species represented in marine, brackish and freshw ater environments with pelagic, epibenthic 
and infaunal habitats (Read 2007). Although bris tleworm lengths range from 1 mm to 3 m, 
the majority are less than 10 cm (Ruppert et al. 2004). Bristleworms are differentiated by the 
presence of paired appendages (parapodia) on mo st segments that contain bundles of bristles 
(chaetae).  
 Bristleworms have varied life histories and reproductive strategi es (Gherardi et al. 2007, Read 
2007). Although the majority of species repr oduce sexually, asexual reproduction does occur 
(Ruppert et al. 2004, Bybee et al. 2007, Gherardi et al. 2007). Fe rtilised gametes may develop 
in the plankton, be brooded, or develop with in the adult body (resu lting in live birth) 
(Giangrande et al. 2000, Bybee et al. 2007). Most  bristleworms have subannual life-spans, 
although some species, such as the known invasive Sabella  spallanzanii, may live for more 
than one year (Bybee et al. 2007, Read 2007, Ramey 2008).    Ecologically, bristleworms are im portant as food for other animal s, bioturbators and habitat 
providers (Hutchings 1998, Volkenborn & Reis e 2007). Bristleworms can be divided into 
sedentary and mobile forms. Sedentary forms ar e suspension feeders, living either infaunally 
or epifaunally in tubes, burrows or calcareous tubes (Ruppert et al. 2004). Suspension feeding 
forms commonly have elaborate appendages for f eeding and respiration and their gregarious 
nature and tubes provide habitat for other organisms (Dubois et al. 2002). Mobile forms include active swimmers, crawlers and burrowe rs, and may be carnivores, herbivores or 
scavengers (Ruppert et al. 2004). Bristleworm s can be extremely abundant and may reach 
densities of thousands of individuals per m
2 (Hutchings 1998, Dubois et al. 2002, Essink & 
Dekker 2002, Labrune et al. 2007).  There are many well known occurrences of non- indigenous bristleworms impacting upon core 
values. For example, Polydora  spp. Hydroides ezoensis  and Terabrasabella heterounicinata 
are recognised aquaculture pests, responsible for the loss of significan t quantities of stock, 
such as oysters in Japan, abalone in the US A and scallops in Norway (Culver & Kuris 2000, 
Nishi 2002, Radashevsky & Olivares 2005, Simon et al. 2005). Reef-forming bristleworms, 
such as Ficopomatus enigmaticus, can dramatically alter community and sediment characteristics (Wilson et al. 1996 In Hutchings 1998, Schwindt et al. 2001, Bruschetti et al. 
2008). Some introduced species remain restricted  to disturbed environments such as ports 
where their habitat forming capabilities may in crease biodiversity (R anasinghe et al. 2005, 
Cinar et al. 2006, Marrs undated).     Given that bristleworms: 
 are known components of marine  biofouling assemblages; 
 are known to have been introduced to new locations; 
 have known impacts on core values; 
it is concluded that bristleworm species repr esent a potential hazard via vessel biofouling. 
 
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  61  
4.5.2. Risk Assessment 
4.5.2.1.  Entry assessment 
Bristleworms are pr
esent in the majority of marine environments and are most prominent in 
soft-bottom sediments (Hutchings 1998, Read 2007). Fouling and errant bristleworms are 
components of shallow marine fouling assemb lages worldwide and are likely to be found 
within ports and other vessel hubs (OECD 1967). Latitudinal trends in general bristleworm 
diversity are contradictory and uncertain (Hutchings 1998, Giangrande & Licciano 2004, 
Metcalfe & Glasby 2008). For example, fouling serpulid tubeworms are reported to have 
heavier settlement in tropical waters (O ECD 1967) but this may be confounded by the 
dramatically higher settlement and growth in areas of high nutrients and sheltered conditions, 
such as the Mediterranean, the Indian Ocean, and ports in general (OECD 1967, Simon et al. 
2005, Marrs undated).    
 Bristleworms are common foulers of vessel hulls (Lewis 2004). Fo r example, in the 
MAFBNZ commissioned research on biofoulin g 22% of vessels sampled had bristleworm 
fouling (Inglis et al. in prep.). Tubeworms, in  particular, can be dominant fouling organisms 
aided by their ability to grow  and reproduce rapidly (Lewis 2004), and their tolerance of 
previously established algal growth (OECD 1967). Moreover, a number of tubeworms show 
tolerance towards heavy metals, including coppe r and tributyltin, which aids in both their 
fouling of vessels and survival in port  areas (OECD 1967, Lewis 2004, Lewis et al. 2006, 
Burlinson & Lawrence 2007, Qui et al. 2007). Errant  bristleworms have also been recorded as 
hull biofouling (Coutts & Dodgshun 2007), however , unlike tubeworms errant bristleworms 
require existing biofouling to provide shelter and food.   Bristleworms are known to be translocated via hull fouling and remain viable. Lewis et al. 
(2006) attribute hull fouli ng to the introduction of H. sanctaecrucis  to Australia from the 
Pacific side of the Americas or possibly Asia. The introduction of H. ezoensis  to and 
throughout Europe from the north-west Pacific has also been attributed to hull fouling 
(Zibrowius 1992). Thirty three sp ecies of non-indigenous bristleworms have been recorded in 
New Zealand (Kospartov et al. in press).  Due to their tubes remaining after death, the actual numbers of viable tubeworms entering 
receiving environments is uncertain. However,  although the original organism is gone, the 
tube still represents a potential shelter or substrate for other non-indigenous organisms.   
 Given that bristleworms: 
 are known to occur along vessel pathways; 
 are known to be associated with vessel biofouling; 
 are known to survive translocation; 
it is concluded that the likelihood of entry of  non-indigenous bristlew orm species into New 
Zealand via vessel biofou ling is non-negligible. 
4.5.2.2.  Establishment assessment 
Bristleworms entering as biofouling are likely  to f
ind suitable habitat in New Zealand, 
particularly in ports of first arrival, where there are abundant settleme nt substrates and high 
nutrient levels. Moreover, many bristleworm species can tolera te both polluted and pristine 
habitats (OECD 1967, Simon et al. 2005, Bruschetti  et al. 2008). Errant bristleworms are not 
likely to be restricted by feedi ng specialisation, therefore it is  likely that if an errant 
bristleworm can survive translocation it will  be able to survive in New Zealand port 
environments.   
62  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
Bristleworms are able to transfer from vessels to New Zeala nd substrates through release of 
gametes/larvae. Some mobile species may be ab le to swim but most successful releases will 
be in association with other fouling being removed. Release, settlement and growth rates of 
larval worms may depend on specific temperat ures or conditions (O ECD 1967, Sion et al. 
2005, Bybee et al. 2007). For exampl e, OECD (1967) notes that “ Hydroides norvegica  in the 
mouth of the Clyde took three months to attain  a length which specimens in Madras Harbour 
reached in three weeks.” In temperate waters, reproduction and settlement generally occurs in 
summer (OECD 1967, Currie et al. 2000, Giangrande  et al. 2000, Simon et al. 2005). Ports 
offer sheltered, nutrient rich environments increasing the likelihood of establishment if 
propagules are released (Simon et al. 2005, Marrs undated). As sexual reproduction is 
probably contingent on conspecifi cs being close enough to receive  spawning cues (Ruppert et 
al. 2004), mature conspecifics must reach crit ical density to trigger gamete release. 
 The physical movement of adult non-indige nous bristleworms from vessels onto New 
Zealand substrates may also result in populat ion establishment.  Many sedentary forms of 
bristleworm can regenerate body parts (Ruppert et al. 2004, Gherardi et  al. 2007), therefore, 
abrasion of individuals from th e hull surface may result in viable adults entering the New 
Zealand environment. Alternatively, entire t ubes or aggregations may be dislodged from the 
hull surface. This is particular ly relevant if poor hull cleani ng practices are employed. Errant 
bristleworms can swim off the hull and onto r eceiving substrates where they may establish. 
Spawning requirements will become a factor after the establishment of adult populations/individuals.   There are known recent recordings of non-i ndigenous bristleworms establishing in New 
Zealand. Of the thirty three sp ecies of non-indigenous bristlewo rms have been recorded in 
New Zealand with eighteen confirmed as establis hed (Kospartov et al. in  press). In the last 
five years three species ( H. ezoensis , S. spallanzanii , and Paralepidonotus ampulliferus ) have 
become established (Kospartov et al. in press).   Given that particular bristleworm species: 
 are known to inhabit similar climates/environmen ts to those that exist in New Zealand; 
 are capable of forming establishi ng populations in New Zealand; 
 have been recently recorded as  established in New Zealand; 
it is concluded that the likelihood of establis hment of non-indigenous bristleworm species in 
New Zealand via vessel biof ouling is non-negligible. 
 
4.5.2.3.  Consequence assessment 
The establishm
ent of non-indige nous bristleworms in New Zealand is likely to result in 
impacts to benthic ecosystems. For example, tube dwelling species may produce reefs which 
modify physical habitat or ch ange plankton communities and li ght attenuation through filter-
feeding (Holloway 1999, Bruschetti et al. 2008).  Bristleworms may f acilitate biodiversity 
loss or gain (Holloway 1999) w ith impacts being most apparent when individuals are densely 
aggregated (Ross et al. 2007). For example, the mud tubes built by P. ciliata  can agglomerate 
to form layers of mud up to 20 cm thick. Thes e layers can eliminate the native flora and fauna 
or alter ecosystem dynamics (Daro and Polk 1973 In NIMPIS 2010). By contrast, crab 
populations in the SW Atlantic incr eased following the introduction of F. enigmaticus reefs, 
which then impacted other organisms th rough predation (Schwi ndt et al. 2001).  
 
Some introduced bristleworms have been reported to increase diversity in disturbed/polluted 
environments where they provide habitat and food (Ranasinghe et al . 2005, Cinar et al. 2006, 
Marrs undated). However, br istleworm reefs formed by F. enigmaticus, have been shown to 
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  63  
have had positive effects on the densities of ot her non-indigenous species when compared to 
habitats created by native species (Heiman et  al. 2008). Therefore, th e establishment of non-
indigenous bristleworm reefs may aid the establishment of other non-indigenous species. It is 
noted, however, that bristleworms can be quite  ephemeral and colonies can wax and wane 
with anthropogenic and natural disturban ces (Zulkhe, 2001, Dubois et al. 2002, Chollett & 
Bone 2007).   
 The proliferation of calcareous tubes on vessels  can compromise antifouling systems, block 
intake pipes, increase cleaning costs and decrease fuel effici ency. In addition, bristleworms 
may impact the economic value of aqu aculture and fisheries. For example, Polydora  spp. are 
well known shell-borers that are able to thrive under aquaculture conditions, i.e. nutrient rich 
environments with dense organism aggreg ations. Examples of negative impacts of Polydora  
spp. include: greater than 50% mortality in abal one/paua (Lleonart et al. 2003a), reduced meat 
and shell weight in oysters ( C. gigas ) (Royer et al. 2006), and the loss of one third of 
Norways 1997 scallop spat (Mortensen et al. 2000). In addition, tubeworms, of both boring 
and non-boring species, compete with cultu red bivalves for food and oxygen (Ropert & 
Goulletquer 2000).   In New Zealand,  P. websteri  and P. hoplura are non-indigenous tubeworm species that have 
had negative effects on oyster aquaculture (Handley 1995, Handley & Bergquist 1997). Other 
boring species such as Boccardia knoxi also have similar effects (Llenart et al 2003b). The 
recent incursion response to S. spallanzanii  has cost New Zealand $1 million, with a further 
$1.5 million earmarked for further control.   
4.5.3. Risk Estimation 
Given that particular bristleworm
 species: 
 are commonly associated with vessel biofouling; 
 are capable of forming establishi ng populations in New Zealand; 
 have recently established populations within New Zealand; 
 have well known invasive behaviours; 
 can be ecosystem engineers; 
 can facilitate the entry and es tablishment of a variety of other non-indigenous organisms; 
 are known to impact marine ecosystems; 
 are known to impact aquaculture  and other economic values; 
it is concluded that the impor t of non-indigenous bristleworm species via biofouling poses a 
non-negligible risk to New Zealand’s marine core values. 
 
4.5.4. Risk Management 
A generic, b
road spectrum management opti on for biofouling should mitigate bristleworm 
associated risks (Section 3).  However, the mobility of bristleworms presents further 
management difficulties, and as a result, systems that contain all organisms in-water prior to 
treatment may be required to redu ce the risk of pre-treatment es cape. In addition, the ability of 
bristleworms to regenerate body parts and the po tential to spawn should be considered during 
the application of management options.  
64  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
4.5.5. References 
Bruschetti, M; Luppi, T
; Fanjul , E; Rosenthal, A; Iribarne, O (2008) Grazing effect of the 
invasive reef-forming polychaete Ficopomatus enigmaticus (Fauvel) on phytoplankton 
biomass in a SW Atlantic coastal lagoon. Journal of Experime ntal Marine Biology  345: 
212-219. 
Burlinson, F C; Lawrence, A J (2007) A comparis on of acute and chronic toxicity tests used 
to examine the temporal stability of  a gradient in copper tolerance of Hediste 
diversicolor  from the Fal estuary, Cornwall, UK. Marine Pollution Bulletin  54: 66-71.  
Bybee, D R; Bailey-Brock, J H;  Tamaru, C S (2007) Gametogene sis and spawning periodicity 
in the fan worm Sabellastarte spectablis  (Polychaeta: Sabellidae). Marine Biology 151: 
639-648. 
Chollett, I; Bone, D (2007) Effects of hea vy rainfall on polychates: Differential spatial 
patterns generated by a large scale disturbance. Journal of Experimental Biology and 
Ecology 340: 113-125. 
Cinar, M E; Katagn, T; Oztur k, B; Egemen, O; Ergen, Z; Ko catas, A; Onen, M; Kirkim, 
F; Bakir, K; Kurt, G; Dag li, E; Kaymakci, A; Acik, S; Dogan, A; Ozcan, T (2006) 
Temporal changes of soft- bottom zoobe nthic communities in and around Alsancak 
Harbour (Izmir Bay, Aegean Sea), with specia l attention to the au tecology of exotic 
species. Marine Ecology 27: 229-246. 
Coutts, A D M; Dodgshun, T J (2007) The nature a nd extent of organism s in sea-vessels: a 
protected mechanism for marine bioinvasions. Marine Pollution Bulletin  54: 875-886. 
Culver, C S; Kuris, A M (2000) The apparent er adication of a locally established introduced 
marine pest. Biological Invasions  2: 245-253. 
Dubois, S; Retiere, C; Olivier, F (2002) Biodiversity associated with Sabellaria alveolata  
(Polychaeta: Sabellariidae) reefs:  effects of human disturbances. Journal of the Marine 
Biological Association of the United Kingdom  82: 817-826. 
Essink, K; Dekker, R (2002) Gene ral patterns in invasion ecol ogy tested in the Dutch Wadden 
Sea: the case of a brackish- marine polychaetous worm. Biological Invasions  4: 359-368. 
Gherardi, M; Sciscioli, M; Lepore, E; Todisco, G; Giangr ande, A (2007) The role of 
schiogenesis in population dynamics of Timarete filigera  (Polychaeta: Cirratulidae) 2-
years observations in the Port of Bari (South Adriatic Sea). Marine Ecology 28: 1-9. 
Giangrande, A; Licciano, M (2004)  Factors influencing latitudina l pattern of biodiversity: an 
example using Sabellidae (Annelida, Polychaeta). Biodiversity and Conservation 13: 
1633-1646. 
Giangrande, A; Licciano, M; Pagliara, P; Gambi, M C (2000)  Gametogenesis and larval 
development in Sabella spallanzanii  (Polychaeta: Sabellidae )  from the Mediterranean 
Sea. Marine Biology  136: 847-861. 
Handley, S J (1995) Spionid polyc haetes in Pacific Oysters, Crassostrea gigas  (Thunberg) 
from Admiralty Bay, Marlborough Sounds, New Zealand. New Zealand Journal of 
Marine and Freshwater Research  29: 305-309. 
Handley, S J; Bergquist, P R ( 1997) Spionid polychaete infestat ions of intertidal Pacific 
oysters Crassostrea gigas  (Thunberg), Maharangi Harbour, northern New Zealand. 
Aquaculture  153: 191-205. 
Heiman, K W; Vidargas, N; Micehi, F (2008) . Non-native habitat as home for non-native 
species: comparison of communities associat ed with invasive tubeworm and native 
oyster reefs. Aquatic Biology 2: 47-56.  
Holloway, M G (1999) Ecological eff ects of the Introduced Fanworm, Sabella spallanzanii, 
on hard substrate epifaunal assemblages.  PhD Thesis; University of Melbourne, 
Australia. 
Hutchings, P (1998) Biodiversity and functioni ng of polychaetes in benthic sediments. 
Biodiversity and Conservation  7: 1133-1145. 
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  65  
Inglis, G J; Floerl, O; Unwi n, M; Ponder-Sutton, A; Seawar d, K; Kospartov, M; Bell, A; 
Kluza, D (in prep.). The biosecurity risks associated with biofouling on international 
vessels arriving in New Zealand. Summary of the patterns and predictors of fouling.  
Biosecurity New Zealand Technical Paper. 
Kospartov K; Inglis G; Seawar d K; Brink A van den; D’Arch ino R; Ahyong S (in press) Non-
indigenous and cryptogenic mari ne species in New Zealand . Current state of knowledge 
– Interim report prepared for MAFBNZ Post Border Directorate. MAF Biosecurity New 
Zealand. Wellington, New Zealand. 
Labrune, C; Gremare, A; Amouroux, J-M; Sarda,  R; Taboada, S (2007) Assessment of soft-
bottom polychaete assemblages in the Gulf  of Lions (NW Mediterranean) based on a 
mesoscale survey. Estuarine, Coastal and Shelf Science  71: 133-147. 
Lewis, J (2004) Hull fouling as a vector for the translo cation of marine organisms: Report 1: 
Hull fouling research . Ballast Water Research Se ries Report No. 14. Dept of 
Agriculture, Fisheries and Fore stry; Canberra, Au stralia; 129 pp. 
Lewis, J A; Watson, C; ten Hove, H A (2006) Establishment of the Caribbean serpulid 
tubeworm Hydroides sanctaecrucis  Kroyer [in] Morch, 1863, in Northern Australia. 
Biological Invasions 8: 665-671. 
Marrs, R (undated) Invasive non-native species in the UK. Hydroides ezoensis . 
http://138.253.199.114/IAAP%20Web/IAAPwebs ite/Marinesppintro2.asp?ID=30. 
Accessed 20 February 2009. 
Metcalfe, K N; Glasby, C J (2008) Diversity of Polychaeta (Annelida) and other worm taxa in 
mangrove habitats of Darwin Harbour, northern Australia. Journal of Sea Research  59: 
70-82.  
NIMPIS (2010). Polydora ciliata  impact details. National Introduced Marine Pest 
Information System (Eds: Hewitt C.L., Martin R.B., Sliwa C., McEnnulty, F.R., 
Murphy, N.E., Jones T. & Cooper, S.). Web publication 
<http://crimp.marine.csiro.au/nimpis>, Accessed: 2nd February 2010. 
Nishi, E (2002) Hydroides ezoensis. In: The Ecological Society of Japan (Eds) Handbook of 
Alien Species in Japan. Chijin-Shokan, Tokyo; p180. (in Japanese) 
Qiu, J-W; Tang, X; Zheng, C; Li, Y; Huang, Y (2007) Copper complexation by fulvic acid 
affects copper toxicity to the larvae of the polychaete Hydroides elegans . Marine 
Environment Research  64: 563-573. 
Radashevsky, V I; Olivares, C (2005) Polydora uncinata  (Polychaeta: Spionidae) in Chile: an 
accidental transportation across the Pacific. Biological Invasions  7: 489-496. 
Railkin, A I (2004) Marine Biofouling: Colonization Processes and Defences. CRC Press; 
303pp. 
Ramey, P (2008) Life history of a dominant polychaete, Polygordius jouinae, in inner 
continental shelf sands of the Mid-Atlantic Bight, USA. Marine Biology  154: 443-452. 
Ranasinghe, J A; Mikel, T K; Velarde, R G; Weisberg, S B; Montagne, D E;  Cadien, D B; 
Dalkey, A (2005) The prevalence of non-indigenous species in southern Californian 
embayments and their effects on benthic macroinvertebrate communities. Biological 
Invasions  7: 679-686. 
Read, G (2007) Bristleworms (Class Polychaeta of  Phylum Annelida). In MacDiarmid A (Ed) 
The Treasures of the Sea: Nga Toanga a T angaroa. A Summary of the Biodiversity in 
the New Zealand Marine Ecoregion . WWF-New Zealand; We llington, New Zealand. 
Ross, D J; Keough, M J; Longmore, A R; Kno tt, N A (2007) Impacts of two introduced 
suspension feeders in Port Phillip Bay, Australia. Marine Ecology Progress Series  340: 
41-53. 
Ruppert, E E; Fox, R S; Barnes, R D (2004) Invertebrate Zoology: A functional evolutionary 
approach (7th edition). Thompson Lear ning; Victoria, Australia. 
Schwindt, E; Bortolus, A; Iribarn e, O O (2001) Invasion of reef- builder polychaet e: direct and 
indirect impacts on the na tive community structure. Biological Invasions 3: 137-149. 
66  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  67 Simon, C A; Kaiser, H; Britz, P J (2005) The life history responses of the abalone pest, 
Terebrasabella heterouncinata , under natural and aquaculture conditions. Marine 
Biology 147: 135-144. 
Volkenborne, N; Reise, K (2007) Effects of Arenicola marina  on polychaete functional 
diversity revealed by large-scale experimental lugworm exclusion. Journal of Sea 
Research  57: 78-88.  
Zibrowius H (1992) Ongoing modification of th e Mediterranean marine and flora by the 
establishment of exotic species. Mesogee 51: 83-107. 
Zuhlke, R (2001) Polychaete tubes create epheme ral community patterns: Lanice conchilega 
(Pallas, 1766) associations studied over six years. Journal of Sea Research 46 (3-4): 
261-272. 
    
4.6. BRYOZOANS (PHYLUM: BRYOZOA) 
4.6.1. Hazard Identification 
Bryozoans are typically sessile colonial m
arine organisms of wh ich there are over 4,000 
known species. Bryozoans are common components of shallow marine an d continental shelf 
biota with a few species found in deep sea a nd freshwater environments (Gordon & Mawatari 
1992, Hughes 2001, Stone 2006). Although most bryozoans are sessile colonial benthic filter 
feeders, one genus is solitary and motile and on e polar species forms pelagic colonies (Peck et 
al. 1995, Ruppert et al. 2004). Representatives can be found in environments ranging from 
tropical to polar and sheltered to energetic  (Barnes & Whittington 1999, Creary 2002, Ruppert 
et al. 2004, Mallela 2007).    The typical bryozoan life history involves a short pelagic larval  phase and a sessile colonial 
adult phase. Larvae are brooded be fore being expelled into the water column for minutes to 
hours, before settling on nearby substrate. A si ngle settled larva transforms into the adult 
form, called a zooid, which is composed of th e animal encased in a tiny calcareous tube 
(Ruppert et al. 2004). This first zooid is calle d the ancestrula. A bryoz oan colony is formed 
following consecutive budding of zooids from th e ancestrula and then from the other zooids. 
After colony formation many zooids exhibit pol ymorphisms that assist with the division of 
labour, for example, the zooids on the edge of the colony continue to bud while others may be 
solely dedicated to cleaning the colony surf ace or reproduction (Ryland 1965). Although all 
colonies have ovaries and testes, the developm ent and release of eggs and sperm is usually 
staggered to allow enough cross- fertilisation between colonies to ensure outbreeding (Ruppert 
et al. 2004).  Bryozoans are ecologically important as reef builders and as a result create habitat for a 
variety of organisms (Martindale 1992, Stone 2006, Mallela 2007, Gordon et al. 2009). 
Bryozoans are able to dominate benthic substrates in energetic environments, with encrusting forms growing quickly to occupy available space  and smothering other organisms. In more 
sheltered environments erect forms out-compete  other organisms by vertical growth (Ward & 
Thorpe 1989, Barnes & Whittingt on 1999, Gordon et al. 2009). 
 Bryozoa are known components of marine biof ouling assemblages and have a history of 
introductions (Ryland 1965, Gordan & Mawata ri 1992). For example, the introduction of 
Tricellaria inopinata  to the Venice Lagoon, Italy, resulted in almost total dominance of the 
lagoon’s benthic habitat between 1989 and 1993 and reduced the native bryozoan fauna to small marginal habitats (Occhipinti Ambrogi  2000). Non-indigenous bryozoa may modify 
habitat and provide hard substrates for othe r organisms including other non-indigenous 
species (Zabin et al. 2007).   Given that bryozoans: 
 are known components of marine  biofouling assemblages;  
 are known to have been introduced to new locations; 
 have known impacts on core values; 
it is concluded that bryozoan species repres ent a potential hazard via vessel biofouling. 
 
68  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
4.6.2. Risk Assessment 
4.6.2.1.  Entry assessment 
Bryozoans are found worldwide although diversity is richest at m
id to high latitudes, with an 
emphasis in the southern hemisphere (Barnes & Griffith 2008). Australasia is considered the 
centre of bryozoan biodiversity, with New Zeal and currently considered the richest in 
diversity with the highest per centage of endemic species (60% ) (Barnes & Griffith 2008). As 
common members of fouling commun ities in ports it is likely th at propagules will encounter 
vessels in port, especially within Aust ralasia (Ryland 1965). 
 
While bryozoans have multiple possible entry pathways, including floating debris, other organisms and ballast water (Mackie et al. 2006 ), hull fouling is probably the most common 
introduction pathway (Key et al. 1999, Barnes  & Fraser 2003, Barnes et al. 2004, Barnes & 
Milner 2005). Generally bryozoans are one of the more common hull fouling organisms, and 
are often amongst the first and most rapid co lonisers of hull surfaces (Ryland 1965, Railkin 
2004). Many bryozoans are tolerant  of the heavy metal concentra tions found in ports and anti-
fouling paints and are attracted to the bare metal of the rudder shaft and propeller (Wisely 
1963, Piola & Johnston 2006a). They are also know n to grow profusely in pipes (Railkin 
2004).   Typically, hull fouling bryozoans are those that inhabit estuaries and ports (Ryland 1965, Gordan & Mawatari 1992). As a result, these sp ecies are likely to have broad tolerances to 
changing temperature, salinity and turbidit y. For example, Allen (1953) noted that 
Watersiporia cucullata , found on many Australian Naval vessels, survived voyages of up to 
three months at speeds reaching 30 knots. Moreover, bryozoans, particularly the seasonal species, can regress or become “dormant” dur ing unfavourable conditi ons and rejuvenate 
once conditions become favourable, a behaviou r which may aid in translocation (Keough 
1986, Piola & Johnstone 2006b). In terms of morphology, both encrusting (e.g. W. 
subtorquata ) and branching species (e.g. Bugula neritina) are known to survive translocation 
as hull fouling (Ryland 1965).    All non-indigenous bryozoans introduced to New Zealand are likely to have been introduced via hull fouling (Kospartov et al. in press) . In the MAFBNZ commissioned research on 
biofouling, bryozoans were found on 26% of vessels  entering New Zealand (Inglis et al. in 
prep.). Inglis et al. (in pr ep.) noted that 17 of the 30 non-indigenous bryozoan species 
translocated to New Zeala nd were not yet established. 
 Given that particular bryozoan species: 
 are known to occur along vessel pathways; 
 are known to be associated with vessel biofouling; 
 are known to survive translocation; 
it is concluded that the likelihood of entry of non-indigenous bryozoa n species into New 
Zealand via vessel biofou ling is non-negligible. 
 
4.6.2.2.  Establishment assessment 
Bryozoans are likely to find suitable habitat and environm
ents in New Zealand. The majority 
of hull fouling bryozoans are likely to be limited only by the temperature or environment 
encountered in New Zealand’s ports of first arrival. Low temperatures are known to affect 
bryozoan growth, reproduction, and mortality (Amui-Vedal et al. 2007, Borque et al. 2007). As a result, at-risk locations to temperate br yozoa establishment are ports and marinas during 
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  69  
spring and summer months. Due to low temperatur es, tropical bryozoa are unlikely to be able 
to establish in New Zealand.    
 Bryozoans are able to transfer from vessels th rough the release of gametes, larvae or colony 
fragments. Bryozoan growth a nd reproduction are generally great est in the summer months, 
which coincide with the period of high water productivity in New Z ealand, making this time 
the most likely for bryozoan establishment (Ryland 1965, Amui-Vedal et al. 2007, Borque et al. 2007). Opportunities for propagule release whic h will result in successful establishment 
are probably restricted to port and marina environments. Moreover, given the short time 
bryozoan larvae spend in the water column and the short distance viable fragments are likely 
to disperse, these environments are the most  likely point of initial establishment.   
 There are recent recordings of  non-indigenous bryozoa establis hing in New Zealand. Of the 
23 established non-indigenous species of bryozoa in New Zealand five have been reported 
since 2000 (Kospartov et al. in press). Moreover,  42% of bryozoans were identified as non-
indigenous in a survey of New Zeal and ports (Gordon & Mawatari 1992). 
 Given that particular bryozoan species: 
 are known to inhabit similar climates/environmen ts to those that exist in New Zealand; 
 are capable of forming establishi ng populations in New Zealand; 
 have been recently recorded as  established in New Zealand; 
it is concluded that the likelihood of estab lishment of non-indigenous bryozoan species in 
New Zealand via vessel biof ouling is non-negligible. 
 
4.6.2.3.  Consequence assessment 
The consequences of non-indigenous bryozoan establishm
ent to New Zealand may range 
from nuisance fouling to modi fied habitat and ecosystem f unction. Introduced bryozoa in 
New Zealand have already been found to sign ificantly impact environmental and economic 
values. For example, W. arcuata overgrew all subtidal benthic habitat in the Auckland region 
when first introduced in 1957 (Skerman 1960, Mort on and Miller 1968). Mo re recently, in the 
same region and habitat, Zoobotryon verticillatum  has formed gelatinous masses up to a metre 
long (Gordon et al. 2009). Similarly, Membraniporopsis tubigera has negatively impacted 
fisheries as large calcareous formations have  fouled fishing equipment (Gordon et al. 2006). 
 
Bryozoa introductions have also impacted va lues elsewhere. For example, the large 
calcareous formations formed by Schizoporella variabilis  in the soft sediments of San 
Francisco Bay provide habitat for a range of organisms (Zabin et al. 2007). This has had both positive and negative consequences as while th e overall biodiversity has increased the 
bryozoans provide habitat for a number of non- indigenous species, for example, they are a 
‘nursery’ for the commercially destructive Atlantic oyster drill, Urosalpinx cinera (Zabin et 
al. 2007). In addition, in the Gulf of Maine, USA, colonies of the non-indigenous bryozoan 
Membranipora membranacea have formed on kelp fronds leaving them brittle and susceptible to breakage. As the kelp forest has recede d, the habitat has been taken over by the non-
indigenous algae Codium fragile  ssp. tomentosoides (Harris & Tyrrell 2001, Pratt & Grason 
2007).   Although it is unlikely that hull maintenance wi ll have to be undertaken on account of 
bryozoan colonies alone, it must be considered that dead bryozoan colonies facilitate the 
attachment and growth of other fouling organisms including calcareous worms, hydroids and those that do not readily  settle on anti-fouling paint (Allen 1953, Floerl  et al. 2004, Floerl et 
al. 2005).  
70  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
 
Consequences to biodiversity  and habitats may be restri cted to disturbed and port 
environments. Hull fouling and port dwelling br yozoan’s tolerance to  pollution and thermal 
variation may not give them an advantage outside of these environments (Gordon & Mawatari 1992, Piola & Johnston 2006).   
4.6.3. Risk Estimation 
Given that particular bryozoan species: 
 have recent records of entry in to New Zealand via hull fouling; 
 have recent records of establishm
ent in New Zealand; 
 have well known invasive behaviours; 
 may facilitate the entry and establishment of a variety of non-indigenous organisms; 
 may have significant impacts on core values, such as, those attri buted to aquaculture 
and biodiversity; 
it is concluded that the impor t of non-indigenous bryozoan species via biofouling poses a non-
negligible risk to New Zeal and’s marine core values. 
 4.6.4. Risk Management 
The generic broad spec 
trum approach to biofouling mitigation is considered appropriate for 
mitigation of the risks associated with hull fouling bryozoans (Section 3). Specific 
considerations for hull fouling bryozoans should include appropriate local containment due to 
the likelihood of release of associated non-indigenous organisms into the receiving 
environment.  
4.6.5. References 
Allen, F E (1953) Distribution of Marine Invertebrates by Ships. Australian Journal of 
Marine and Freshwater Research  4: 307-316. 
Am
ui-Vedel, A-M; Hayward, P J; Porter, J S (2007) Zooid size and growth rate of the 
bryozoan Cryptosula pallasiana  Moll in relation to temperature, in culture and its 
natural environment. Journal of Experimental Marine Biology and Ecology 353: 1-12. 
Barnes, D K A; Griffiths, H J (2008) Biodiversity and biogeog raphy of southern temperate 
and polar bryozoans. Global Ecology and Biogeography  17: 84-99. 
Barnes, D K A;  Fraser, K P P (2003) Rafti ng by five phyla on man-made flotsam in the 
Southern Ocean. Marine Ecology Progress Series  262: 289-291. 
Barnes, D K A; Milner, P (2005) Drifting plastic and its conse quences for sessile organism 
dispersal in the Atlantic Ocean. Marine Biology  146: 815-825. 
Barnes, D K A; Whittington, M (1999) Biomechanics and mass mortality of erect bryozoans 
on a coral reef. Journal of the Marine Biological Association of the United Kingdom  79: 
745-747. 
Barnes D K A; Warren N L; We bb K; Phalan B; Reid K (20 04) Polar pedunculate barnacles 
piggy-back on pycnogona, penguins, pi nniped seals and plastics. Marine Ecology 
Progress Series  284: 305-310.  
Borque, D; Davidson, J; MacNair, N G; Arsena ult, G; LeBlanc, A R; Landry, T; Miron, G 
(2007) Reproduction and early life hist ory of an invasive ascidian Styela clava  Herdman 
in Prince Edward Island, Canada. Journal of Experimental Marine Biology and Ecology  
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Cocito, S; Sgorbini, S; Bi anchi, C N (1998) Aspects of  the biology of the bryozoan Pentapora 
fascialis  in the northwestern Mediterranean. Marine Biology 131: 73-82. 
Creary, M M (2002) Spatial dist ribution of epibenthic byozoans found on the roots of 
Rhizophora mangle , Kingston Harbour, Jamaica, W.I. Bulletin of Marine Science  73(2): 
477-490. 
Floerl, O; Inglis, G J (2003) Boat ha rbour design can exace rbate hull fouling. Austral Ecology  
28: 116-127. 
Floerl, O; Pool, T K; Inglis, G J (2004) Posi tive interactions betw een non-indigenous species 
facilitate transport by human vectors. Ecological Applications  14: 1724-1736. 
Floerl, O; Inglis, G J; Marsh, H M (2005) Selectivity in vector  management: an investigation 
of the effectiveness of measures used to prevent transport of non-indigenous species. 
Biological Invasions  7: 459-475. 
Gordon, D P; Mawatari, S F (1992) Atlas of ma rine-fouling bryozoa of New Zealand ports 
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Gordon, D; Smith, F (2007) Lace Corals (Phyl um Bryozoa). In MacDiarmid, A (Ed) The 
Treasures of the Sea: Nga Taonga a Tangaroa. A Summary of the Biodiversity in the 
New Zealand Marine Ecoregion.  WWF-New Zealand; Wellington, New Zealand. 
Gordon, D P; Ramalho, L V; Tayl or, P D (2006) An unreported invasive bryozoan that can 
affect livelihoods – Membraniporopsis tubiger in New Zealand and Brasil. Bulletin of 
Marine Science  78(2): 331-342. 
Gordon, D P; Taylor, P D; Bi gey, F P (2009) Phylum Bryozoa. In Gordon, D P (ed) New 
Zealand Inventory of Biodiversity. Volume 1 : Kingdom Animalia; Radiata, 
Lophotrochozoa, Deuterostomia. Canterbury University Press; Christchurch, New 
Zealand; pp 271-297. 
Harris, L G; Tyrrell, M C (2001) Changing community  states in the Gulf of Maine: synergism 
between  invaders, overfishing and climate change. Biological Invasions  3: 9-21. 
Hayward, B; Morley, M (2009) Contrasts in th e recent introductions of two tunicates to 
northern New Zealand. In O’Shea, S; Costello, M J; O’Connell, K (eds) Programme 
and Abstracts for Annual Conference in Auckland 2-4 September 2009 . Accessed at 
http://nzmss.rsnz.org, p 92. 
Hughes, D J (2001) Quantitative analysis of a deep-water bryozoan collection from the 
Hebridean continental slope. Journal of the Marine Biologi cal Association of the United 
Kingdom . 81: 987-993. 
Inglis, G J; Floerl, O; Unwi n, M; Ponder-Sutton, A; Seawar d, K; Kospartov, M; Bell, A; 
Kluza, D (in prep.). The biosecurity risks associated with biofouling on international 
vessels arriving in New Zealand. Summary of the patterns and predictors of fouling.  
Biosecurity New Zealand Technical Paper. 
Keough, M J (1986) The distribution of the br yozoan Bugula neritina on seagrass blades: 
settlementgrowth and mortality. Ecology 67: 846-857. 
Key, M M Jr; Winston, J E; Volpe, J W; Jeffrie s, W B; Voris, H K (1999) Bryozoan fouling 
of the Blue Crab Callinectes sapidus  at Beaufort, North Caro lina. Bulletin of Marine 
Science 64(3): 513-533.  
Kospartov K; Inglis G; Seawar d K; Brink A van den; D’Arch ino R; Ahyong S (in press) Non-
indigenous and cryptogenic mari ne species in New Zealand . Current state of knowledge 
– Interim report prepared for MAFBNZ Post Border Directorate. MAF Biosecurity New 
Zealand. Wellington, New Zealand. 
Lewis, J (2004) Hull fouling as a vector for the translo cation of marine organisms: Report 1: 
Hull fouling research . Ballast Water Research Se ries Report No. 14. Dept of 
Agriculture, Fisheries and Fore stry; Canberra, Au stralia; 129 pp. 
Lewis, P N; Hewitt, C L; Riddl e, M; McMinn, A (2003) Marine introductions in the Southern 
Ocean: an unrecognised hazard to biosecurity. Marine Pollution Bulletin  46: 213-223. 
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Lewis, P N; Riddle, M J; Hewitt, C L (2004) Management of exogenous threats to Antarctica 
and the sub-Antarctic Islands: balancing risk s from TBT and non-indi genous organisms. 
Marine Pollution Bulletin  49: 999-1005. 
Mackie, J A; Keough, M J; Christidis, L (2006) Invasion patterns inferr ed from cytochrome 
oxidase I sequences in  three bryozoans, Bugula neritina, Watersiporia subtorquata , and 
Watersiporia arcuata. Marine Biology 149: 285-295. 
Mallela, J (2007) Coral reef encruster commun ities and carbonate production in cryptic and 
exposed coral reef habitats along a gradient of terrestrial disturbance. Coral Reefs 26: 
775-785.  
Martindale, W (1992) Calcified ep ibionts as palaeological tools: examples from Recent and 
Pleistocene reefs of Barbados. Coral Reefs  11: 167-177. 
McCann, L D; Hitchcock, N G; Winston, J E; Ru iz, G M (2007) Non-native bryozoans in 
coastal embayments  of th e Southern United States: New records for the western 
Atlantic. Bulletin of Marine Science  80(2): 319-342. 
Morton, J; Miller, M (1968) The New Zealand Seashore . Collins; Auckland, New Zealand; 
638 pp. 
Occhpinti Ambrogi, A (2000) Biotic Invasions in a Mediterranean Lagoon. Biological 
Invasions  2: 165-176. 
O’Dea, A (2003) Seasonality and zooid variation in Panamanian encrusting bryozoans. 
Journal of the Marine Biologica l Assosiation of the United Kingdom 83: 1107-1108. 
Peck, L S; Hayward, P J; Spencer-Jones, M E (1995) A pelagic bryozoan from Antarctica. 
Marine Biology 123:757-762. 
Piola, R F; Johnston, E L (2006a) Differential tolerance to metals among populations of the 
introduced bryozoan Bugula neritina. Marine Biology 148: 997-1010. 
Piola, R F; Johnstone, E L (2006b)  Differential resistance to exte nded copper exposure in four 
introduced bryozoans. Marine Ecology and Progress Series  311: 103-114. 
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http://www.bowdoin.edu/~mpratt/invasives.html - accessed 8 July 2008. 
Pratt, M C; Grason, E W (2007) Invasive species as a new food source: does a nudibranch 
predator prefer eating an invasive bryozoan? Biological Invasions  9: 645-655. 
Railkin, A I (2004) Marine Biofouling: Colonization Processes and Defences. CRC Press; 
303pp. 
Ryland, J S (1965) Catalogue of Main Fouling Organisms (found on ships coming into 
European water): Polyzoa. Organisation for Economic Co-operation and Development; 
Paris, France. 
Rubin, J A (1985) Mortality and avoidance of competitive overgrowth in encrusting Bryozoa. 
Marine Ecology Progress Series 23: 291-299. 
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approach (7th edition). Thompson Learning; Victoria, Australia. 
Saunders, M; Mexetas, A (2008) High recruitmen t of the introduced bryozoan Membranipora 
membranacea is associated with kelp bed defoliation in Nova Scotia, Canada. Marine Ecology Progress Series 369: 139-151. 
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scale sp
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74  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis Wood, R A (1999) Reef Evolution . Oxford University Press. 
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Cambridge, Massachusetts. 
 
  
4.7. CRABS (INFRAORDER: BRACHYURA) 
4.7.1. Hazard Identification 
The Brachy
ura, or true crabs, are successf ul and highly specialised decapods. This group 
contains over 4,500 species, the majority of which are found in benthic marine habitats, 
however, freshwater and terres trial species exist (Ruppert et  al. 2004). Crabs range in size 
from only a few millimetres wide to the Japanese  spider crab whose chelipeds may span three 
metres (Ruppert et al. 2004). Although the majori ty of crabs cannot swim, notable exceptions 
are members of the Portunidae fa mily (Ruppert et al. 2004).   
 
Ecologically, crabs are important components of benthic communities due to their feeding and 
foraging activities. Many crabs are active food ga therers and feed upon a range of organisms, 
for example, the highly invasive Carcinus maenus  has been observed to prey upon organisms 
from more than 100 families and 158 genera in  5 plant/protist and 14 animal groups (Cohen et 
al. 1995). Predation by crabs can regulate the abundance and species composition of local 
fauna, such as, marine and estuarine infa unal soft-bottom communities (Quammen 1984), 
epibenthic microbiota (Foreman 1985) and me iofauna (Hoffman et al. 1984). Burrowing and 
foraging activity can also impact the environmen t, especially in estuarine intertidal areas 
(Montague 1982, DePatra & Levin 1989). For exampl e, crabs can retain  nitrogen in the 
system, increase oxygenation and drainage of the soil (Montague  1982), and increase 
meiofauna (DePatra & Levin 1989). In addition, crabs provide food for many apex predators 
such as sharks and octopuses and are targeted by humans for consumption both commercially and recreationally.  The life-cycle of crabs involves a pelagic larval stage and a mob ile benthic adult stage. Crabs 
have separate sexes and usually copulate, ex changing spermatophores, with the eggs brooded 
by the females (Ruppert et al. 2004). The majority  of crabs spawn in se awater irrespective of 
whether they inhabit freshwater  or terrestrial environments (Ruppert et al. 2004). Brackish-
water forms, such as blue crabs ( Callinectes  spp.) and river immigrants such as Eriocheir 
sinensis, return to more saline waters for breeding and larval release (Rud nick et al. 2005).  
 Species of crab are known to have establishe d outside their native ra nge and some have had 
severe ecological, economic and social impacts. C. maenas, Charybdis japonica  and 
E. sinensis  are well known as invasive marine organi sms. These crabs, like many high impact 
invasive marine organisms, ar e highly fecund, broadly tolerant  to abiotic va riation, active 
food gatherers and can impact core values. For example C. maenus , E. sinensis , Hemigrapsus 
sanguineus  and Rithropanopeus harrisii  have all been recorded f eeding upon fish trapped in 
nets, bait fish and fish and mussels subject to aquaculture (Cohen et al. 1995, Brousseau et al. 2001, Rudnick et al. 2005, Roche & Torchi n 2007). The monetary impact of E. sinensis  in 
Germany since 1912 is estimated at approximately 80 million Euro. This estimation included impacts to catchment gear installation and maintenance, bank erosion, feeding, and on 
commercial fisheries and aquaculture (Gollasch 2006).   Given that crabs: 
 are known components of marine  biofouling assem
 blages;  
 are known to have been introduced to new locations; 
 have known impacts on core values; 
it is concluded that crab species represen t a potential hazard via vessel biofouling. 
 
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  75  
4.7.2. Risk Assessment 
4.7.2.1.  Entry assessment 
Crabs have a global distribution with diversity decreasing with  inc
reasing latitude (Bruce & 
Patuawa 2007). The southwest Pacific is specifically noted for high diversity in many groups 
(Bruce & Patuawa 2007). In addi tion to shelter, ports and harbours provide a productive 
habitat for crabs as water temperatures are often higher than the open coast (Cohen et al. 
1995). Both larval and adult life-stages are, ther efore, likely to encounter vessels within ports 
and marinas.     Most crab introductions ar e attributed to ballast water, dr y ballast, and aquaculture (Cohen et 
al. 1995, Carlton & Cohen 2003, Her borget al. 2005, Miller et al  2006, de Lafontaine et al. 
2008) however, crabs have also be en found in association with hu ll fouling (Farrapiera et al. 
2007) and within seachests (Coutts et al. 2003, Coutts & Dodgshun 2007). For example, Minchin and Gollasch (2003)  observed hundreds crabs, Porcellana longicornis,  within the 
shells of dead barnacles on a commercial vessel  hull. In addition, the Gl obal Invasive Species 
Database ( http://www.issg.org/database/welcome/
) fact sheet on E. sinensis  notes that “some 
specimens in empty shells of cirripeds have been reported on a ships hull.” Pre-moult and moulting crabs also have a tendency to seek shel ter which may result in their transport within 
vessel crevices (Dodgs hun & Coutts 2003).  
 Despite the instances recorded above, crabs ar e rarely found in hull f ouling surveys and are 
usually low in abundance. This may be due to ad ult and larger juvenile crabs requiring a high 
level of fouling to provide ad equate food and shelter, however , heavily fouled vessels are 
generally rare. The available sheltered space, such as dead barnacle shells and gratings, may also limit the number and size of crabs.   Crabs are known to survive translocation with th e capacity to tolerate high variation in abiotic 
conditions being a common attribute among introduced crabs. For example, adult C. maenus 
can withstand over wintering temperatures as low as 0
oC and temperatures as high as 31-36oC 
for brief periods (Hidalgo et al. 2005). C. maenus are also capable of resisting salinities as low 
as 4 ppt (Hidalgo et al. 2005).   Cuculescu et al. (1998) observed th at eurythermal species of crab  are able to survive larger 
changes in temperatures than stenothermal species. In addition, crabs sampled in summer 
survived exposure to higher temperatures than those sampled in winter. This indicates that 
individuals translocated in summer would be able to survive higher temperature changes in 
transit. Larvae and juvenile crab s are more sensitive to temper ature and salinity differences 
than adult crabs (Baylon and Suzuki 2007).   There are recent known introductions of non- indigenous crabs to New Zealand. In the 
MAFBNZ commissioned research on biofouling, Inglis et al. (in prep .) found 7 species of 
non-indigenous crabs that are not established in New Zealand.   
 Given that particular crab species: 
 are known to occur along vessel pathways; 
 are known to be associated with vessel biofouling (although usually restricted to 
heavily fouled vessels); 
 are known to survive translocation; 
it is concluded that the likelihood of entry of  non-indigenous crab spec ies into New Zealand 
via vessel biofouling is non-negligible.  
76  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  77 4.7.2.2.  Establishment assessment 
There are known suitable habita ts for introduced crabs in Ne w Zealand. The sim
ilarity of 
various receiving environments (i.e. climate a nd physical habitat) has been suggested as a 
major factor contributing to th e success of invasive crab species (Lohrer et al. 2000). 
Although many invasive crabs have high tolerance to various abiotic conditions, they are most 
likely to establish in similar non-indigenous habitats to their native environments (Lohrer et 
al. 2000, Thresher et al. 2003, R udnick et al. 2005, Gust & Inglis 2006, Roche & Torchin 
2007). Resources play a significant part in th e success of non-indigenous species, whether 
through superior competition or utilisation of  a free resource (Lohrer et al. 2000, Kado 2003). 
Therefore, the depauperate nature and absence of large coastal crabs in New Zealand (Dell 
1968, Morton 2004, Bruce & Patuawa 2007) may present open niches that will aid the 
establishment of non-indigenous crabs. As crab diversity generally decr eases with increasing 
latitude (Bruce and Patuawa 2007), northern New Zealand is more likely to host suitable 
environments for the establishment of non-indigenous crabs.   Crabs can transfer from a vesse l through release of e ggs, larvae or individual migration into 
the environment where the vessel is moored/b erthed. However, reproduction is required to 
establish a self sustaining population, for exam ple, there are many cases of individual crabs 
being recorded beyond their natura l ranges that have not result ed in established populations 
(Carlton & Cohen 2003, Thresher et al. 2003). Crabs tend to ha ve localised recruitment 
(Thresher et al. 2003, Roman & Palumbi 2004) a nd spread (Thresher et al. 2003, Herborg et 
al. 2005), therefore, reproductiv e success is likely if propa gule numbers are high enough. 
Depending on larval settlement, a self-sustain ing population with low genetic variation may 
also form.     Temperature and salinity are likely to play a cr ucial role in whether larvae released into the 
water column develop, settle, and contribute towards establishment. In general, higher 
temperatures
11 result in faster larval developmen t (Epifanio et al. 1998, Rudnick et al. 2005, 
Baylon & Suzuki 2007). However, the megalope stage of development requires a particularly narrow temperature and salinity range (Epifa nio et al. 1998, Carlton & Cohen 2003, Bravo et 
al. 2007). The tolerance of adu lt crabs to temperature and salinity variation increases 
dramatically after larval sett lement, however, temperature is important for the regulation of 
adult growth, development and reproduction. For example, in southern China E. sinensis 
becomes sexually mature within 1-2 years (Jin  et al. 2001, Zhang et al. 2001), by comparison 
the same species in the San Francisco estuary and in northern Europe can take up to 3 and 5 
years to mature, respectively (Gollasch 1999, Rudnick et al. 2005). Temperature can be 
protective of non-indigenous es tablishment, for example, C. maenas  has been found in 
numerous tropical locations such as Sri Lanka and Brazil, but ha s not been able to establish 
(Carlton & Cohen 2003). Further, if temperatures are particularly low, crab growth and reproduction may be slowed or ev en halted until the next seas on (Rudnick et al. 2005). As a 
result of the influence of temperature on the crab life-cycle, the most conducive location for 
the establishment of non-indigenous crabs is  likely to be northern New Zealand.  
 There are recent known examples of non-indige nous crabs establishing in New Zealand. 
Kospartov et al. (in press) list  four species that have esta blished since 2000, however, it is 
uncertain whether these species we re translocated via hull fouling. 
 The likelihood of establishment appears to  be dependent on many factors. Though an 
organism may be introduced to a suitable e nvironment, establishment depends on a raft of 
conditions being favourable, such as climate and oceanography. This is exemplified by the 
                                                
 
11 As a general rule, the summer temperatures experienced within the crabs native range.  
fact that despite being introduced to Engl and on several occasions and being already 
established on the European continent, E. sinensis  did not establish populations in England 
until recently (Herborg et al. 2005 ). Human mediated transport has increased the distribution 
of non-indigenous crabs, for example, the wide distribution of E. sinesis  and the long distance 
transport and establishment of C. maenas  and C. japonica  populations (Anon 2001 In Gust 
and Inglis 2006, Thresher et al. 2003, Carlton & Cohen 2003). Wh ile crabs are likely have a 
low frequency on vessel pathways, as international maritime traffic intensifies (Hewitt et al. 
2004), the increased vessel dockings will result in increases to propagule exposure and thus 
the likelihood of establishment (Roman  & Palumbi 2004, Petersen 2006, Roman 2006, 
Herborg et al. 2007).  Population numbers of invasive crabs can be highly variable. Under normal conditions crab 
recruitment is localised, however, given the right conditions populations can occasionally 
explode. Environmental conditions are sometimes  conducive to the retention of the large 
larval output in the immediate and surrounding areas, for example, Gimenez & Dick (2007) 
found that the numbers of C. maenas settled in the North Sea ar e related to the direction, 
strength and duration of the wind. In the Tham es (UK) and San Francisco (USA) estuaries 
high recruitment of invasive crabs was observed during dry winters, wh en freshwater input 
and flow-through was lower (Carlton & Cohe n 1995, Herborg et al. 2005). In addition, 
C. maenas and R. harrisii  both spread further up the wester n North American coast when the 
El Nino current advected onto the coast and swept northwards (Petersen 2006). However, 
after these climatically driven  events, populations can dwindle and be restricted to small 
pockets. In contrast to population explosions, populations can also experience rapid declines 
due to unfavourable conditions, for exampl e, the cooler Auckland summer of 2002-03 
resulted in a 66% dr op in catch rates of C. japonica (Gust & Inglis 2006). 
 
Given that particular crab species: 
 are known to inhabit similar climates/environmen ts to those that exist in New Zealand; 
 are capable of forming establishi ng populations in New Zealand; 
 have been recently recorded as  established in New Zealand; 
it is concluded that the likelihood of establis hment of non-indigenous crab species in New 
Zealand via vessel biofou ling is non-negligible. 
 
4.7.2.3.  Consequence assessment 
Crabs can be active predators im
pacting upon commercial and recreational fishing. Native 
New Zealand crabs are not particularly large, ag gressive or wide-sprea d in marine near-shore 
environments, however, the majority of invasive crab species have the aforementioned 
characteristics (Gust & Inglis 2006, MacDonald et al. 2007). As many invasive crabs are 
known to actively and aggressively seek out food sources, commercial  and recreational 
fisheries may be impacted, for example, inva sive crabs may monopolise baited traps used to 
commercially catch New Zealand paddle crabs ( Ovalipes catharus). They may also become 
stuck in fishing gear and feed on bait or  catch reducing fishing efficiency.     
 The introduction of the European green crab ( C. maenus) to North America resulted in severe 
impacts on shore communities, affecting the su rvival and recruitmen t of gastropods and 
bivalves (Vermeij 1982b; Williams 1984 In NIMPIS 2010). The introduction of this crab was 
one of the factors respon sible for the decline of the soft shelled clam ( Mya arenaria ) fishery 
(Whitlow et al. 2003 In NIMPIS 2010), which affected the livelihoods of thousands of 
workers in eastern North America. At the time th e crab extended its range into the fisheries 
areas, catches fell from 14.5 million pounds in 1938 to 2.3 million pounds in 1959 (IUCN 
2009). C. maenus has subsequently establ ished in the west coast of North America, South 
78  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
Africa, south eastern Australia, Brazil, Panama, the Red Sea, Pakistan, Sri Lanka, Myanmar, 
Japan, Patagonia and Hawai’i.  In Australia (Tasmania) C. maenas  has been reported to be a 
significant predator of the clam species, Katelysia scalarina  (Walton et al. 2002 In NIMPIS 
2010). In addition, since the arrival of C. maenus to California in 1993, clam and native shore 
crab populations have dropped significantly (NIMPIS 2010).  The introduction of non-indigenous species may not create immediate imp acts to core values, 
as founding populations may require a shift in th eir new range to encounter a more suitable 
habitat. For example, th e giant red king crab ( Paralithodes camtschaticus), introduced by 
Russia to the Barents Sea in 1960 to form a fishery did not establish numbers nor have 
recorded impacts for approximate ly twenty years. In the 1980’s this crab had migrated to 
Norway and by the 1990’s had increased to num bers significant enough to raise concerns 
amongst local fishermen and environmentalists (IUCN 2009). It has since been shown that 
P. camtschaticus has impacted on local scallop populat ions (Jorgensen 2005). In Norway, 
these crabs are capable of foraging on 150 to 335 g of scallop per m
2 over 48 hr compared to 
the natural biomass of the commercially harvested I celand scallop ( Chlamys islandica ) of 400 
to 1200 g per m2 (Jorgensen 2005).  
 Social and cultural values may also be impacted by established non-indigenous crab 
populations. In addition to affecti ng employment related to fisherie s, the aggressive nature of 
the crabs may make them a nuisance for recreati onal users of coastal waters, such as beach 
users or fishermen trying to clean fishing g ear. Invasive crabs may feed on local bivalve 
species which make up an important commercial and food component of M āori life and for 
recreational fishers.   Established non-indigenous crabs may modify bi otic communities and physical environments. 
The burrowing behaviour of crabs may impact marine shorelines and riverine banks by 
altering erosion rates and sediment characteri stics of nearby and down- current areas. Possible 
knock-on effects may include the smothering of be nthic organisms and reduced water quality.   
 
4.7.3. Risk Estimation 
Given that particular crab species: 
 are asso 
ciated with vessel biofouling (when extensive fouling communities are present); 
 are capable of forming establishi ng populations in New Zealand; 
 have recently established populations with in New Zealand (although, the vector remains 
unknown); 
 have well known invasive behaviours; 
 are known to impact core values; 
it is concluded that the impor t of non-indigenous crab spec ies via biofouling poses a non-
negligible risk to New Zeal and’s marine core values. 
 
4.7.4. Risk Management 
The cryptic nature of crabs m
akes their presence  on vessel hulls difficult to predict or detect. 
Although a hull may appear clean, due to habita t preferences, crabs may be located behind 
gratings that have macro-fouling or reside wi thin the tests of large or small barnacles. 
Therefore, given their association with biof ouling, generic measures are likely to be 
appropriate (Section 3). However, the mobility  of crabs presents further management 
difficulties, and as a result, systems that contai n all organisms in-water prior to treatment may 
be required to reduce the risk  of pre-treatment escape. 
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  79  
4.7.5. References 
Baylon, J; S
uzuki, H (2007) Effects of changes in salinity and temperature on the survival and 
development of larvae and juve niles of the crucifix crab Charybdis feriatus  (Crustacea: 
Decapoda: Portunidae). Aquaculture  269: 390-401. 
Bravo, M A; Cameron, B; Metaxas,  A (2007) Salinity tolerance in the early larval stages of 
Carcinus maenas  (Decapoda, Brachyura), a recent inva der of the Bras d’Or Lakes, 
Nova Scotia, Canada. Crustaceana 80: 475-490. 
Brousseau, D J; Filipowicz, A; Baglivo, J A (2001)  Laboratory investigations on the effects of 
predator sex and size on prey selection by the Asian crab, Hemigrapsus sanguineus . 
Journal of Experimental Marine Biology and Ecology 262(2): 199-210. 
Bruce, N; Patuawa, P (2007) Crustaceans (S ubphylum Crustacea). In MacDiarmid, A (Ed) 
The Treasures of the Sea: Nga Taonga a T angaroa. A Summary of the Biodiversity in 
the New Zealand Marine Ecoregion . WWF-New Zealand; We llington, New Zealand. 
Carlton, J T, Cohen, A N (2003) Episodic gl obal dispersal in shallow water marine 
organisms: the case history of the European shore crabs Carcinus maenas  and C. 
aestuarii . Journal of Biogeography 30: 1809-1820. 
Clark, P F; Rainbow, P S; Robbi ns, R S; Smith, B; Yeomans, W E; Thomas, M; Dobson, G 
(1998) The alien Chinese mitten crab,  Eriocheir sinensis  (Crustacea: Decapoda: 
Brachyura), in the Thames catchment. Journal of the Marine Bi ological Association of 
the United Kingdom  78: 1215-1221. 
Cohen, A N; Carlton, J T; Fountain, M C (1995) Introduction, dispersal and potential impacts 
of the green crab Carcinus maenas  in San Francisco Bay, California. Marine Biology 
122: 225-237. 
Coutts, A D M; Dodgshun, T J (2007) The nature and extent of organisms in seachests: a 
protected mechanism for marine bioinvasion. Marine Pollution Buletin  54: 875-886.  
Coutts, A D M; Moore, K M; Hewitt, C L (2003)  Ships’ seachests: an overlooked transfer 
mechanism for non-indigenous species? Marine Pollution Bulletin  46: 1504-1515. 
Cuculescu, M; Hyde, D; Bowler, K (1998) Therma l tolerance of two species of marine crab, 
Cancer pagurus  and Carcinus maenas . Journal of Thermal Biology  23: 107-110. 
Dell, R K (1968) Composition and distributi on of the New Zealand Brachyuran fauna. 
Transactions of the Royal Society of New Zealand 10: 225-240. 
DePatra, K D; Levin, L A (1989) Evidence of th e passive deposition of meiofauna into fiddler 
crab burrows. Journal of Experimental Marine Biology and Ecology  125(3): 173-192. 
Epifianio, C E; Dittel, A I; Park, S; Schwal m, S; Fouts, A (1998) Early life history of 
Hemigrapsus sanguineus , a non-indigenous crab in the Mi ddle Atlantic Blight (USA). 
Marine Ecology  Progress Series  170: 231-238. 
Foreman, K (1985) Do predators, resources or physical factors control the seasonal cycle of 
meiofauna of marshes? Estuaries 8: 48. 
Gimenez, L; Dick, S (2007) Settlement of shore crab Carcinus maenas  on a m esotidal open 
habitat as a function of transport mechanisms. Marine Ecology Progress Series  338: 
159-168. 
Gollasch, S (2006) NOBANIS – Invasive  Alien Species Fact Sheet – Eriochier sinensis . 
www.nobanis.org. Accessed 9 Jan 2009. 
Gust, N; Inglis, G J (2006) Adaptive multi-scal e sampling to determine an invasive crab’s 
habitat usage and range in New Zealand. Biological Invasions  8: 339-353. 
Hanfling, B; Carvalho, G R; Brandl R (2002) mt-DNA seque nces and possible invasion 
pathways of the Chinese mitten crab. Marine Ecology Progress Series  238: 307-310. 
Herborg, L-F; Weetman, D; Oosterhout, O V; Hanfling, B (2007) Genetic population 
structure and contemporary dispersal patterns of a recent European invader, the Chinese 
mitten crab, Eriocheir sinensis . Molecular Ecology  16: 231-242. 
80  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
Herborg, L M; Rushton, S P; Clare, A S; Bentley, M G (2005)  The invasion of the Chinese 
mitten crab ( Eriocheir sinensis ) in the United Kingdom and its comparison to 
continental Europe. Biological Invasions  7: 959-968. 
Hewitt, C L; Campbell, M L; Thresher, R E; Ma rtin, R B; Boyd, S; Cohen, B F; Currie, D R; 
Gomon, M F; Keough, M J; Lewis, J A; Lockett, M M; Mays, N; MacArthur, M A; 
O’Hara, T D; Poore, G C B; Ross, D J; Storey, M J; Watson, J E; Wilson, R S (2004) 
Introduced and cryptogenic species in Po rt Phillip Bay, Victoria, Australia.  Marine 
Biology 144: 183-202. 
Hidalgo, F J; Baron, P J; Orensanz, J M (2005)  A prediction comes true: the green crab 
invades the Patagonian coast. Biological Invasions  7: 547-552.  
Hoffman, J A; Katz, J; Bertne ss, M D (1984) Fiddler crab deposit-feeding and meifaunal 
abundance in salt marsh habitats. Journal of Experimental Marine Biology and Ecology 
82: 161-174. 
Inglis, G J; Floerl, O; Unwi n, M; Ponder-Sutton, A; Seawar d, K; Kospartov, M; Bell, A; 
Kluza, D (in prep.). The biosecurity risks associated with biofouling on international 
vessels arriving in New Zealand. Summary of the patterns and predictors of fouling.  
Biosecurity New Zealand Technical Paper. 
IUCN (2009) Marine menace: alien invasive  species in the marine environment. 
http://www.iucn.org/about/work/programmes / marine/marine_resources/marine_publicat
ions/?3255/marinemenace . Accessed 17 November 2009. 
Jin, G; Li, Z; Xie, P (2001) The growth patt erns of juvenile and precocious Chinese mitten 
crabs, Eriocheir sinensis  (Decapoda, Grapsida), stocked in  freshwater lakes of China. 
Crustaceana  74: 261-273. 
Jorgensen, L L (2005) Impact scenario for an introduced decapod on Arctic epibenthic 
communities. Biological Invasions  7: 949-957. 
Kado, R (2003) Invasion of Japanese shores by the NE Pacific barnacle Balanus glandula  and 
its ecological and biogeographical impact. Marine Ecology Progress Series . 249: 199-
206. 
Ko, H S (2005) First zoea of Eriphea sebana (Crustacea: Decapoda: Xa nthoidea: Menippidae) 
hatched in the laboratory. Integrative Biosciences 9: 135-138. 
de Lafontaine, Y; Sevigny, J-M; Calve, R; Verreault, G; Despatie, S-P; Veilleux, E (2008) 
Chinese mitten crabs (Eriocheir sinensis ) in the St. Lawrence River and Estuary, 
Canada: new records and risk of invasion. Aquatic Invasions  3(2): 153-163. 
Lohrer, A M; Whitlatch, R B; Wada, K; Fukui , Y (2000) Home and Away: comparisons of 
resource utilization by a marine speci es in native and invaded habitats. Biological 
Invasions  2: 41-57. 
MacDonald, J A; Roudez, R; Glover, T; Weis , J S (2007) The invasive green crab and 
Japanese shore crab: behavioural interactions with a native crab species, the blue crab. 
Biological Invasions  9: 837-848.   
Maeda, M; Itami, T; Furumo to, A; Hennig, O; Imamura, T;  Kondo, M; Hirono, I; Aoki, T; 
Takahashi, Y (1998) Detection of penaei d red-shaped DNA virus (PRDV) in wild-
caught shrimp and other crustaceans. Fish Pathology 33: 381-387. 
McDermott, J J (1998) The western Pacific brachyuran ( Hemigrapsus sanguineus : 
Grapsidae), in its new habitat along the Atla ntic coast of the United States: geographic 
distribution and ecology. ICES Journal of Marine Science  55: 289-298. 
Minchin, D; Gollasch, S (2003) Fouling and sh ips’ hulls: how changing circumstances and 
spawning events may result in th e spread of exotic species. Biofouling  19 (supplement): 
111-122. 
Montague, C L (1982) The influence of fiddl er crab burrows and burrowing on metabolic 
processes in saltmarsh sediment. In Kennedy, V S (Ed) Estuarine Comparisons. 
Academic Press; New York, USA; pp 238-301. 
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82  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis Morton, J (2004) Seashore Ecology of New Zeal and and the Pacific. David Bateman Ltd; 
Auckland, New Zealand; 504 pp.  
NIMPIS (2010). Carcinus maenas  impact details. National Introduced Marine Pest 
Information System (Eds: Hewitt C.L., Martin R.B., Sliwa C., McEnnulty, F.R., 
Murphy, N.E., Jones T. & Cooper, S.). Web publication <http://crimp.marine.csiro.au/nimpis>, Accessed: 2
nd February 2010. 
Pardo, M L; Palma, A T; Priet o, C; Sepulveda, P; Valdivia, I;  Ojeda, P (2007) Processes 
regulating early post-settlement  habitat use in a subtidal  assemblage of brachyuran 
decapods. Journal of Experimental Marine Biology and Ecology 344: 10-22. 
Payen, G G; Bonami, J R (1979) Mise en évidence de particules d'allure virale associées aux 
noyaux des cellules mésodermiques de la z one germinative testiculaire du crabe 
Rhithropanopeus harrisii  (Gould) (Brachyura, Xanthidae). Revue des Travaux de 
l'Institut des Pêches Maritimes  43: 361-365. 
Petersen, C (2006) Range expansion in the nort heast Pacific by an estuary mud crab – a 
molecular study. Biological Invasions  8: 565-576. 
Quammen, M L (1984) Predation by shorebirds, fish, and crabs on invertebrates in intertidal 
mudflats: an experimental test. Ecology 65(2): 529-537. 
Reyns, N B; Eggleston, D B (2004)  Environmentally-controlled, density-dependant secondary 
dispersal in a local estuarine crab population. Oecologia  140: 280-288. 
Roche, D G; Torchin, M E (2007) Established pop ulation of the North American Harris mud 
crab, Rhithropanopeus harrisii  (Gould 1841) (Crustacea: Brac hyura: Xanthidae) inn the 
Panama Canal. Aquatic Invasions  3: 155-161. 
Rodriguez, G; Suarez, H (2001) Anthropogenic Dispersal of Decapod Cr ustaceans in Aquatic 
Environments. Interciencia  26: 282-288. 
Roman, J (2006) Diluting the founder effect: cr yptic invasions expand a marine invader’s 
range. Proceedings of the Royal Society Series B  273: 2453-2459. 
Roman, J; Palumbi, S R (2004) A global invader at home: populat ion structure of the green 
crab, Carcinus maenas , in Europe. Molecular Ecology  13: 2891-2898. 
Rudnick, D; Veldhuizen, T; Tullis, R; Culver, C; Heib, K; Tsukimura, B (2005) A life history 
model for the San Francisco Estuary population of the Chinese mitten crab, Eriocheir 
sinensis  (Decapoda: Grapsoidea). Biological Invasions  7: 333-350. 
Ruppert, E E; Fox, R S; Barnes, R D (2004) Invertebrate Zoology: A functional evolutionary 
approach (7th edition). Thompson Lear ning; Victoria, Australia. 
Thresher, R; Proctor, C; Ruiz, G; Gurney, R; MacKinnon, C; Walton, W; Rodriguez, L; Bax, 
N (2003) Invasion dynamics of the European shore crab, Carcinus maenas , in Australia. 
Marine Biology 142: 867-876. 
Zhang, T; Li, Z; Cui, Y (2001) Survival, growt h, sex ratio, and maturity of the Chinese mitten 
crab ( Eriocheir sinensis ) reared in a Chinese pond. Journal of Freshwater Ecology 16: 
633-640. 
    
4.8. ECHINODERMS  (PHYLUM: ECHINODERMATA) 
4.8.1. Hazard Identification 
Approxim
ately 6,000 species of echinoderms cover organisms, such as, sea stars 
(Asteroids), brittle stars (Ophiuroids), sea urchins, sand dollars (Echinoids), sea 
cucumbers (Holothurians), and sea lilies a nd feather stars (Crinoids) (Ruppert et al. 
2004).  New Zealand has 618 described echinode rm species covering all five classes, 
of which 37% are endemic to  the EEZ (Gordon et al. 2007). 
 Echinoderms are found in environments ranging from intertidal to abyssal depths and perform ecological functions ranging from pr edator to detritivor e.  Echinoderms can 
be conspicuous components of the bent hic fauna such as the sea urchin ( Evechinus 
chloroticus ) grazing fronts of New Zealand’s rocky reefs (Andrew 1988) or the 
feeding swarms of the sea star, Asterias rubens, in the Irish Sea (Dare 1982).  The 
habitats of sea urchins alternate between species rich kelp forests and sea urchin-dominated “barrens”, these states can alte rnate depending on factors such as storms, 
disease and recruitment (Pearse 2006). So me sea stars are keystone predators 
important for the structure and function of marine benthic communities. Their 
capacity for population outbreaks and signi ficant impacts to areas of fishery and 
aquaculture importance is well documented (Sloan 1980, Menge 1982).   The life-cycle of echinoderms is comparativ ely similar throughout al l the classes. The 
majority of species have separate sexes and are broadcast spawners (Ruppert et al. 
2004). With regard to asexual reproduction, echinoderms have varying abilities to 
regenerate lost body parts and in some cases  two individuals can be produced. Certain 
sea stars can regenerate from as little as one arm and a fifth of the central disc (Ruppert et al. 2004).  Asexual reproduction in the sea star Coscinasterias acutispina  
appears to be governed by a variety of factors including nutrition, temperature and parasitism (Haramoto et al. 2007).  There are few reports on the occurren ce of non-indigenous echinoderms and even 
fewer reports of impacts from translocated echinoderms. However, this may be due to 
a deficiency in the sampling of seachests a nd impact studies.  The northern Pacific sea 
star (Asterias amurensis ), the most high profile invasive echinoderm, has been linked 
to the decline of large bivalves within its  introduced range of Tasmania (Ross et al. 
2002, 2003). Other non-indigenous echinoderms have been recorded within Port Phillip Bay, Australia (Hewitt et al. 2004), a nd along the coast of Israel (Galil 2007).  
 Given that echinoderms: 
 are known components of marine  biofouling assemblages;  
 are known to have been introduced to new locations; 
 have known impacts on core values; 
it is concluded that echinoderm species represent a potential hazard via vessel 
biofouling. 
 
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  83  
4.8.2. Risk Assessment 
4.8.2.1.  Entry Assessment 
Echinoderm
s are generally thought to be tr anslocated beyond their natural range via 
ballast water (Rainer 1995), how ever, there is increasing evidence that vessel internal 
seawater systems (including seachests) can  also be a significant translocation 
mechanism (Thresher 2000 In Ross et al. 2003). For example, Coutts & Dodgshun 
(2007) found a sea cucumber ( unidentified), sea urchin (E. chloroticus ), and the 
eleven armed sea star (Coscinasterias muricata ), all indigenous, within the seachests 
of commercial vessels sampled in Ne w Zealand. In addition, a sea star ( Pateriella 
regularis ) was found amongst substantial algal fouling on a barge situated in the 
Derwent estuary, Australia (Lewis et al. 2006).  Despite the presence of A. amurensis  
being detected in samples of biofouling from vessel hulls us ing molecular probes 
(Hayes et al. 2004), there is no evidence that  echinoderms would be able to remain on 
a clean vessel hull when underway, therefore, their translocation is  probably restricted 
to sheltered niche areas (Thresher 2000 In Ross et al. 2003, Inglis et al. in prep.). 
 
Minimal information exists regarding the e nvironmental tole rances of echinoderms, as 
such there is considerable uncertainty regarding their ability to survive translocation across oceanic boundaries. However, it is known that A. amurensis  survived 
translocation from the temperate regions of the north Pacific to Tasmania and that the 
non-indigenous echinoderms found in Israel were translocated from the Red Sea and 
surrounding areas (Galil 2007). In addition, the eight species (five brittle stars, one sea cucumber, two sea urchins) of non-indigenous echinoderms most recently introduced to New Zealand were all native to the I ndo-Pacific region (Kospartov et al. 2008). As 
a wide variety of echinoderms inhabit high stress environments such as intertidal 
rockpools, it can be assumed that some species will have a range of thermal and physical tolerances. 
 
In addition to the uncertainty regardi ng environmental tolerances, there is no 
knowledge on the effects of anti-fouling pain ts on the various groups of echinoderms. 
However, sea stars (A. rubens) have previously been observed to inhabit environments of high heavy metal contamination (Temara et al. 1998).  Given that particular echinoderm species: 
 are known to occur along vessel pathways; 
 are known to be associated with vessel biofouling; 
 are known to survive translocation; 
it is concluded that the likelihood of entr y of non-indigenous echinoderm species into 
New Zealand via vessel biof ouling is non-negligible. 
4.8.2.2.  Establishment Assessment 
It is assum
ed that New Zealand would pr ovide a suitable environment/habitat for a 
number of non-indigenous echinoderm species. Difficulty in establishment within a 
new environment may depend on the organi sms’ feeding strategy, for example, 
detritivores and omnivores may experience less difficulty meeting dietary requirements than predators and herbivores. This may be due to uncertainty introduced by the differences between native and local fauna, for example, herbivorous echinoderms may experience difficulty recognising unfamiliar algae 
(Prince & LeBlanc 1992) or receive less nut ritional value (Lyons & Schiebling 2007). 
84  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
Alternatively, introduced echi noderms may find readily available and nutritious food 
sources, experience predator/competition release, and go on to become a major 
structuring force in the ecosystem, as has been the case with A. amurensis  in 
Tasmania (Ross et al. 2002, 2003). Translocated from Tasmania to Port Phillip Bay, Victoria (Australia) in 1995, the A. amurensis  population had reached an estimated 75 
million individuals by 2000 (Parry & Cohen 2001).  For establishment in New Zealand waters, echinoderms of both sexes must be present 
within spawning distance. It is assumed that  it is possible for two individuals to be 
present within the one niche area (i.e. seach est) and for spawning to occur within this 
area. Alternatively, individuals may leave th e vessel and encounter individuals of the 
opposite sex in the introduced environment.  At present, there are no known establishmen ts of non-indigenous echinoderms in New 
Zealand (Kospartov et al. in press).  Given that particular echinoderm species: 
 are known to inhabit similar climates/env ironments to those that exist in New 
Zealand; 
 are capable of forming establishi ng populations in New Zealand; 
it is concluded that the likelihood of es tablishment of non-indigenous echinoderm 
species in New Zealand via vesse l biofouling is non-negligible. 
4.8.2.3.  Consequence Assessment 
The potential consequen
ces associated with  large predatory sea stars and sea urchins 
are profound as many are noted keystone species. Sea urchins have the ability to clear macroalgal stands to produce ‘urchi n barrens’ (Pearse 2006), therefore, the 
establishment of non-indigenous sea urchins may exacerbate this phenomenon in the New Zealand environment. In addition, non-indigenous sea urchins would compete 
with indigenous grazing species, some of which have important commercial, 
recreational and cultural fisheries, such as abalon e and kina (Pearse 2006). 
 As predatory sea stars are important to the structure of their na tive ecosystems, their 
introduction and establishmen t may have significant eff ects on bivalve populations. 
For example, the introduction of A. amurensis  to the Derwent Estuary in 1986 is 
likely responsible for the decline and s ubsequent rarity of, surface or shallow 
sediment dwelling, bivalve species (Ross et al. 2003). If other ar eas of the Tasmanian 
coast were to attain the sea star densi ties within the Derwent Estuary, the authors 
predict large direct effects on native assemblages, particul arly on populations of large 
surface dwelling bivalves, including several commercial species. In addition, A. amurensis  is potentially preventing the re-e stablishment and recovery of adult 
populations in the Derwent Estuary, as the diet  of the sea star is found to have a high 
prevalence of juvenile molluscs (Ross et al . 2003). For example, predation losses of 
commercial scallop spat ove r a settlement season are reported to be up to 50% 
(Crawford pers. comm. In Hutson et al. 2005). It has since been shown that the 
Australian scallops ( Pecten fumatus and Chlamys asperrima ) have a low escape 
response frequency to A. amurensis  compared to the indigenous C. muricata  
suggesting that the risk of predation on scal lops will be greater for the introduced sea 
star. Similar effects can be expected for New Zealand bivalve species not previously 
exposed to predatory sea stars. 
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  85  
 
In Port Phillip Bay, Victoria (Australia) A. amurensis been observed predating on spat 
and on grow-out ropes (J. Me rcer, G. Parry, pers. comm. In Dommisse & Hough 
2004). However, despite these oc currences, sea star predati on is not considered a big 
threat to stock losses of commercially  grown mussels (Dommisse & Hough 2004). 
This is perhaps due to the fact that in Port Phillip Bay, spat are "resocked" and 
immersed in freshwater before on-growing for the removal of predators (Gunthorpe et 
al. 2001 In Dommisse & Hough 2004). Although such  practice is not followed in 
Tasmania, those farmers interviewed did not consider stock losses to the sea star as 
significant (Dommisse & Hough 2004). Th e introduction of a non-indigenous 
predatory sea star to New Zealand may th erefore have indirect effects on New 
Zealand aquaculture due to th e alteration of standard prac tices, rather than direct 
effects on aquaculture species. There is li ttle doubt that benthic assemblages have 
been altered where populations of A. amurensis have established. It is likely that the 
introduction of a predatory sea star to  New Zealand would impact on benthic 
assemblages and specific commercial bivalve species. 
 
Apart from those consequences associat ed with predatory sea stars, another 
consequence of non-indigenous echinoderm in troductions is likely to be competition 
for resources. For instance, the sea star Pateriella regularis  shows intraspecific and 
interspecific competition with the chiton Chiton glaucus on the intertidal cobble fields 
of Wellington, New Zealand (Palmer & Ph illips 2009). These effects on marine 
benthic assemblages may have negative impacts on commercial and foods source for 
Māori culture and recr eational gatherers. 
 
4.8.3. Risk Estimation 
Given that particular echinoderm
 species: 
 are associated with vessel biofouling,  particularly in niches areas; 
 are capable of forming establishi ng populations in New Zealand; 
 have well known invasive behaviours; 
 can be keystone species that structure marine communities through predation or 
herbivory; 
 associated with high impacts, such as Asterias amurensis , are present in 
southeastern Australia; 
 have had high impacts on core values fo llowing establishment in other countries; 
it is concluded that the import of non-i ndigenous echinoderm species via biofouling 
poses a non-negligible risk to Ne w Zealand’s marine core values. 
 
4.8.4. Risk Management 
To m
anage the risks associated with echinoderms a generic appro ach to removal of 
fouling is appropriate (Section 3). As seachests  present a likely habitat of translocated 
echinoderms, this area should be given due attention by inspectors on the dock and 
during haul out and cleaning activities. The ability of some echinoderm species to 
regenerate must also be considered duri ng haul out and cleani ng procedures whereby, 
the entire organism, including jettisone d body parts, should be contained. The 
mobility of some echinoderm species presents further management difficulties, and as 
86  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
a result, systems that contain all organi sms in-water prior to treatment may be 
required to reduce the risk of pre-treatment escape.  
  
4.8.5. References 
Andrew, N L (1988) Ecological asp ects of the comm
on sea urchin, Evechinus 
chloroticus , in northern New Z ealand: a review.  New Zealand Journal of 
Marine and Freshwater Research 22: 415-426. 
Coutts, A D M; Dodgshun, T J (2007) The natu re and extent of organisms in vessel 
seachests: A protected mechanism for marine bioinvasions. Marine Pollution 
Bulletin  54: 875–886. 
Dare, P J (1982) Notes on the swarming behaviour and population density of Asterias 
rubens  L. (Echinodermata: Asteroidea) feeding on the mussel, Mytilus edulis  
L ICES Journal of Marine Science  40: 112-118. 
Dommisse, M; Hough, D (2004) Cont rolling the northern Pacific seastar Asterias 
amurensis  in Australia. Final report for the Australian Government Department of 
the Environment and Heritage. 
Gordon, D; Smith, F; Patuawa, T (2007) Sea stars, sea urchins and sea cucumbers 
(Phylum Echinodermata). In MacDiarmid, A (Ed) The Treasures of the sea, 
ngā taonga a Tangaroa. A summary of biodiversity in the New Zealand 
marine ecoregion.  WWF – New Zealand; Wellington, New Zealand. 
Haramoto, S; Komatsu, M; Yamazaki, Y ( 2007) Patterns of ase xual reproduction in 
the fissiparous seastar Coscinasterias acutispina (Asteroidea: Echinodermata) 
in Japan. Zoological Science  24(11): 1075-1081. 
Hayes K; Sutton, C; Gunasekera, R; Sliwa,  C; Migus, S; McEnnu lty, F; Dunstan, P; 
Green, M; Patil, J (2004) Empirical valid ation: Small vessel translocation of 
key threatening species. Stage I – Asterias amurensis .  Department of the 
Environment and Heritage; Canberra, Australia. 
Hutson, K S; Ross, D J; Day, R W; Ahern, J J (2005) Australian scallops do not 
recognise the introduced predatory seastar Asterias amurensis. Marine Ecology 
Progress Series 298: 305–309 
Inglis, G J; Floerl, O; Unwi n, M; Ponder-Sutton, A; Seaward, K; Kospartov, M; Bell, 
A; Kluza, D (in prep.). The biosecurity risks associ ated with biofouling on 
international vessels arriving in New Zealand. Summary of the patterns and predictors of fouling.  Biosecurity New Zealand Technical Paper. 
Lewis, P N; Bergstrom, D M; Whinam, J (2006) Barging in: a temperate marine 
community travels to the subantarctic. Biological Invasions  8: 787-795. 
Menge, B A (1982) Effects of feedin g on the environment: Asteroidea. In: Janguoux 
M and Lawrence J M (eds) Echinoderm Nutrition , pp 521–554. A.A. Balkema, 
Rotterdam, The Netherlands 
Palmer, S; Phillips, N (2009) The ecological role of an omnivorous seastar ( Patiriella 
regularis ) in intertidal cobble fi elds. In O’Shea, S; Co stello, M J; O’Connell, 
K (eds) Programme and Abstracts for Annual Conference in Auckland 2-4 
September 2009 . New Zealand Marine Scienc es Society; Accessed at 
http://nzmss.rsnz.org
; 1-126 pp. 
Parry, G D; Cohen, B F (2001) The dist ribution, abundance andpopulation dynamics 
of the exotic seastar Asterias amurensis during the first th ree years of its 
invasion of Port Phillip Bay. Report N o. 33, Marine and Freshwater Research 
Institute, Queenscliff. 
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  87  
88  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis Pearse, J S (2006) Ecological role of purple sea urchins. Science , 314: 940-941.  
Rainer, S F (1995) Potential for the introducti on and translocation of  exotic species by 
hull fouling: A preliminary assessment .  Centre for Research on Introduced 
Marine Pests, Technical Report 1. CSIRO; Australia.  
Ross, D J; Johnson, C R; Hewitt, C L (2003)  Variability in the impact of an 
introduced predator ( Asterias amurensis : Asteroidea) on soft-sediment 
assemblages.  Journal of Experimental Marine Biology and Ecology  288: 
257–278. 
Ross, D J; Johnson, C R; Hewitt, C L (2002) Impact of introduced seastars Asterias 
amurensis on survivorship of juvenile commercial bivalves Fulvia 
tenuicostata.   Marine Ecology Progress Series  241: 99-112. 
Ruppert, E E; Fox, R S; Barnes, R D (2004) Invertebrate Zoology: A functional 
evolutionary approach  (7th edition). Thompson Lear ning; Victoria, Australia. 
Sloan NA (1980) Aspects of the feeding biology of asteroids. Oceanography and 
Marine Biology: an Annual Review  18: 57–124 
Talman, S G; Keough, M J (2001) Impact of an exotic clam, Corbula gibba, on the 
commercial scallop Pecten fumatus in Port Phillip Bay, south-east Australia: 
evidence of resource restricted growth in a subtidal environment. Marine 
Ecology Progress Series  221: 135-143. 
Temara, A; Skei, J M; Gillan, D; Wa rnau, M; Jangoux, M; Dubois, P (1998) 
Validation of the asteroid Asterias rubens  (Echinodermata) as a bioindicator 
of spatial and temporal trends of Pb, Cd, and Zn contamination in the field. 
Marine Environmental Research  45(4-5): 341-356. 
4.9. FLATWORMS  (PHYLUM: PLATYHELMINTHES) 
4.9.1. Hazard Identification 
The phylum
 Platyhelminthes includes free-li ving flatworms (Turbellaria), parasitic 
flukes (Trematoda) and tapeworms (Cestoda ) (Species 2000). The classes Trematoda 
and Cestoda are entirely parasitic and spend th e vast proportion of their life-histories 
endosymbiotically, and as such, are exempt  from this risk assessment. Within 
Turbellaria, the order Poly cladida (polyclads) contains  greater than 1,000 described 
marine flatworm species. More than half of these species occur between latitudes 30o 
N and 30o S, with approximately two thirds of these species inhabiting the Indo-
Pacific Region (Rawlinson 2008). Marine fl atworms are generally sensitive to 
temperature, thus many species are restricted  within their latit udinal range (Rawlinson 
2008). The order Polycladida is divided in to two suborders, Cotylea and Acotylea, 
based on the presence or absence of a suck er. The Cotylea (sucker present) are more 
species rich in lower latitudes than the Acotylea (Rawlinson 2008).  
 
All flatworms are hermaphrodites, reproduci ng either through copulation or internal 
fertilisation (Ruppert et al. 2004, Lee 2006). In addition, flatworms have a well developed ability to regenerate. Flatworm numbers appear to peak in the summer and 
early autumn (Skerman 1960, Rivest et al. 1999) which coincides with their breeding 
season in which more than one batch of eggs can be produced (Lee 2006). Predatory flatworms often deposit eggs inside the em pty shells of their prey, whereas other 
species lay gelatinous masses, chains, or pl ates of eggs on hard substrata (Skerman 
1960, Lee 2006). Flatworms exhibit brooding be haviour by lying on their eggs which 
develop directly and hatch live young (Ruppert  et al. 2004). As adult flatworms carry 
toxins, they are rarely preyed upon (Lee 2006).  Flatworms are considered to be important mobile predators of sessile hard-substrate 
marine communities (Newman et al. 2000 In Rawlinson 2008). However, some 
species and juveniles of predaceous species  are herbivorous (Ruppert et al. 2004). As 
an order, flatworms feed on a variety of organisms from protozoa, rotifers, small 
crustaceans and annelids to ascidians, bivalves, crustaceans and gastropods (Newman et al. 1993, Rivest et al. 1999, Ruppert et al. 2004, Lee 2006). However, individual 
species of flatworms are usually specia list feeders that show strong feeding 
preferences in both prey species and size (Skerman 1960, Galleni et al. 1980, 
O’Connor & Newman 2003). Flatworms det ect their prey through chemoreception or 
by water movement (Skerman 1960, O’C onnor & Newman 2003, Lee et al. 2006). 
Feeding methods vary from simply ingesti ng or extruding the stomach directly onto 
the prey to the use of mucous and toxins  for prey subdual (Skerman 1960, Galleni et 
al. 1980, Ruppert et al. 2004, Lee 2006).  
Introductions and impacts of non-indigenous flatworms have been recorded globally. 
Flatworms are known to impact sessile mari ne populations, for example, up to 90% of 
the youngest oysters were lost in some area s of the Mediterranean due to flatworm 
predation (Galleni et al. 1980). Flatworm pr edation of cultivated shellfish has also 
caused major impacts in North America and Japan, whilst France has significant 
control measures in place to stop the introduction of flatworms associated with the importation of Japanese oysters (Galleni et al. 1980). However, the impacts associated 
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  89  
with the introduction and esta blishment of non-indigenous fl atworm species have not 
always been recorded. For example, after initial high densities (19 individuals per 
cm2), numbers of Convoluta convoluta  in North America have fallen (to 3 individuals 
per cm2) with no impacts recorded (Rivest et al. 1999, Byrnes & Whitman 2003). 
Likewise, concerns were raised over impact s to aquaculture and wild oyster stocks 
when the Asian flatworm was found in the Netherlands (Sluys et al. 2005), however, 
no impacts have since been recorded. It s hould be noted, that flatworms are not well 
studied and, as a result, there is uncertainty with regard to  the risks associated with 
their introduction. 
 Given that flatworms: 
 are known components of marine  biofouling assemblages;  
 are known to have been introduced to new locations; 
 have known impacts on core values; 
it is concluded that flatworm species re present a potential hazard via vessel 
biofouling.  
4.9.2. Risk Assessment 
4.9.2.1.   Entry assessment 
Flatworm
s are known to inhabit the vessel hulls and they have  previously been 
recorded on these structures in New Zeal and (Coutts & Dodgshun 2007, Inglis et al. 
in prep.), Germany (Gollash 2002) and Brazil (Farrapeira et al. 2007). Some 
flatworms may be able to stay attached to the surface of a relatively clean hull as they 
possess suckers (Minchin & Gollasch 2003). However, many flatworms are known to 
associate with barnacles, oysters and asci dians which they use for shelter, food and 
sites for their egg masses (Skerman 1960, J ohns et al. 2009). As a result, flatworms 
are more likely to be found amongst a biofou ling assemblage rather than on a clean 
open hull.   
 Flatworms are known to remain viable follo wing translocation. Survival following a 
voyage to New Zealand from Australia or the Pacific Islands is t hought to be feasible 
as flatworms are able to undergo extended periods without feedi ng (Galleni et al. 
1980, O’Connor & Newman 2003). However, give n that flatworms are sensitive to 
differences in water temperature (O’Connor & Newman 2003), the survival of adults 
from or through the Pacific Islands is unlik ely, therefore, temperate Australia remains 
the likely source of viable non-indigenous flatworms.   Uncertainty exists with regard the tran slocation of non-indigenous species due to 
taxonomic resolution, and as a result, the ma jority flatworms pres ent in hull fouling 
are only identified at the phylum or order leve l. Therefore, while it  is almost certain 
that flatworms have the ability to be transported via hull fouling, the number of 
positive identifications of non-indigenous speci es is low and as yet none have been 
recorded in New Zealand.   The translocation of the eggs of non-indigenous  species should also be considered as 
flatworms use the empty shells of their prey (e.g. barnacles) as sites to lay their eggs. 
At present, there are no records of egg masses found on vessels, though this is likely 
to be due to a deficiency in targeted inve stigations. Assuming that it is possible for 
90  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
egg masses to be transported via hull biof ouling, there are questions regarding their 
arrival in a viable state. Although brooding is common amon g flatworms, experiments 
conducted on two flatworm species showed that brooding had little effect on hatching 
success (Johnstone & Lee 2008). As the fl atworms studied are relatively common 
Australian species, the possibility of unattended eggs being transported to New 
Zealand as viable propagules is highlighted. However, Johnstone & Lee (2008) note 
that brooding behaviour ma y have advantages under changing salinities and 
temperatures, reinforcing the uncertainty a ssociated with the transit of unattended 
eggs.   There are few recorded occurrences of flat worms on vessel hulls. This may be due to 
their naturally low numbers, although this may also be exacerbated by sampling bias. As flatworms are small, motile and cryptic, often found within barnacles and bivalves, 
there are a variety of reasons that may contribute to an underestimation of their 
numbers. In addition, flatworms are likely to  be associated with high fouling in low 
flow areas, such as the environment provided by seachests, which are perhaps less well studied than the outer hull surface. Seven intern ational vessels were found 
carrying (unidentified) flatworms duri ng the MAFBNZ commissioned research on 
biofouling (Inglis et al. in prep.). Prior to  this research, no introductions of marine 
flatworms to New Zealand via hull fouling had been recorded.  The effect of anti-fouling paint on flatwo rms is unknown as it is uncertain whether 
their slime would form a pr otective layer from the bioc ides. However, their known 
habitats, such as barnacle shells, are likely to provide such protection.  Given that particular flatworm species: 
 are known to occur along vessel pathways; 
 are known to be associated with vessel biofouling; 
 are known to survive translocation; 
it is concluded that the likelihood of entr y of non-indigenous flat worm species into 
New Zealand via vessel biof ouling is non-negligible. 
 
4.9.2.2.  Establishment assessment 
There are likely to be suitable habitats  for the establishm
ent of non-indigenous 
flatworms in New Zealand as they are commonly found under boulders, in crevices of 
oyster clumps and other cryptic locations. Desp ite prey preference being restricted to 
specific organism groups, flatworms may feed on a variety of species within the target 
group (Skerman 1960, Galleni et al. 1980, O’Connor & Newman 2003). As such, 
flatworms are likely to find suitable prey  in a new location. Furthermore, many 
common biofouling species are present in New Zealand ports.   As flatworms are capable of swimming short di stances, they would be  able to transfer 
directly from vessel hulls to local substrata. Non-i ndigenous flatworms can also 
establish a population in New Zealand through the entry of egg masses, sexual 
reproduction and asexual reproduction (fra gmentation and regeneration of an 
individual or multiple individuals).  There are no records of n on-indigenous marine flatworms establishing in New 
Zealand. However, it has been noted that “a t the moment we are hard-pressed to give 
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  91  
definitive names to any of the several sp ecies of ‘oyster leeches’ that are being 
encountered by aquaculturists” (Diggles et al. 2002 In Johns et al. 2009). Therefore, 
in New Zealand, further taxonomic research is required on flatworms not only for the 
protection of important economic values, such as bivalve aq uaculture, but also 
environmental, social and cultural values. At present, it is unlikely that a non-
indigenous flatworm population established in New Zealand would be able to be 
correctly identified.  
 Given that particular flatworm species: 
 are known to inhabit similar climates/env ironments to those that exist in New 
Zealand; 
 are capable of forming establishi ng populations in New Zealand; 
 are unlikely to be correctly identified following establishment in New Zealand; 
it is concluded that the likelihood of es tablishment of non-indigenous flatworm 
species in New Zealand via vesse l biofouling is non-negligible. 
 
4.9.3.  Consequence assessment 
Flatworm 
s are well known pest organisms imp acting the global aqu aculture of oysters 
and mussels (Galleni et al. 1980). If flatworms that prey  on commercial species of 
bivalve were to establish in New Zealand, the likely impacts would include reductions 
in spat and adult growth. Given that the pr ey of flatworms can include both oysters 
and mussels, and the known impacts to comme rcial stocks in other jurisdictions, the 
potential consequences to New Zealand’s aquaculture industry c ould be wide-ranging. 
In 2008, mussels ( Perna canaliculus) and oysters (C. gigas ) comprised of 77% 
(NZ$204.3 million) and 6% of the value (NZ$16.9 million) of New Zealand’s 
aquaculture exports, respectively (Aquacultu re New Zealand 2009).  Given the value 
of bivalve aquaculture to New Zealand, the possible establishment of novel flatworm species via hull biofouling is a non-neglig ible concern. Additi onal costs would be 
involved if protective measures were intr oduced to reduce the number of flatworms 
on farms, the transfer of flatworm between farms, or their occurrence in stock for sale. Further, the presence of pr edatory flatworms may lead to the formation of market 
barriers from countries where popul ations are not established.  
 The impact of non-indigenous flatworms on wi ld bivalve populations in New Zealand 
is unknown. It is likely that environmental vari ation would restrict their numbers in all 
but the most sheltered environments, as has been demonstrated for the invasive C. 
convoluta  in North America (O’Conner & Newman 2003). While flatworms are 
selective consumers of juven ile sessile organisms, the large numbers of settling larvae 
and juveniles tend to swamp these pred atory effects (Skerman 1960, Byrnes & 
Witman 2003). Therefore, it is likely th at non-indigenous flatworms would have a 
localised chronic (low intens ity but sustained) effect on wild bivalve species, the 
consequences of which are unknown. 
4.9.4. Risk Estimation 
Given that particular flatwor
m species: 
 are known components of hull fouling assemblages;  
92  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
 are known to have been introduced to new locations; 
 are unlikely to be correctly iden tified upon entry into New Zealand; 
 are capable of forming establishi ng populations in New Zealand; 
 are likely to have adverse effects on core  values, such as, bivalve aquaculture and 
coastal benthic communities; 
it is concluded that the import of non-i ndigenous flatworm species via biofouling 
poses a non-negligible risk to Ne w Zealand’s marine core values. 
 
4.9.5. Risk Management 
To m
anage the risks presented by flatworm s via hull fouling, a generic approach to 
removal of biofouling is appropriate (Sect ion 3). In addition, the risk management 
strategy should include the removal of ba rnacle shells, and other macro-fouling, due 
to their role as shelter for flatworm adults and eggs. In addition, the mobility of 
flatworms presents further management difficu lties, as a result, systems that contain 
all organisms in-water prior to treatment may be required to re duce the risk of pre-
treatment escape. As flatworms are sensit ive to temperature and salinity changes, 
these factors may be used in the mitigation of viable transfers (O’ Connor & Newman 2001). 
4.9.6. References 
Aquaculture New Zealand (2009) New Zealand Aquaculture far
m facts (2nd edition). 
http://www.aquaculture.org.nz/aquacultu re-in-nz/industry-overview/overview/ . 
Accessed 20 January 2010. 
Byrnes, J; Witman, J D (2003) Impact assessment of an invasive flatworm, Convoluta 
convolute, in the Southern Gulf of Maine. Journal of Experi mental Marine 
Biology and Ecology 293: 173-191. 
Coutts, A D M; Dodgshun, T J (2007) The natu re and extent of organisms in sea-
vessels: a protected mechanism for marine bioinvasions. Marine Pollution 
Bulletin  54: 875-886. 
Farrapeira, C M R; Melo, A V de O M de; Ba rbosa, D F; Silva, K M E da (2007) Ship 
Hull Fouling in the Port of Recife, Pernambuco. Brazilian Journal of 
Oceanography  55(3): 207- 221. 
Galleni, L; Tongiorgi, P; Ferre ro, E; Salghetti, U (1980) Stylochus mediterraneus 
(Turbellaria: Polycladida), predator on the mussel Mytilus galloprovincialis . 
Marine Biology 55: 317-326. 
Gollasch, S (2002) The importance of ship hull fouling as a v ector of species 
introductions into the North Sea. Biofouling 18(2): 105-121. 
Hewitt, C L; Campbell, M L; Thresher, R E; Martin, R B (1999) Marine Biological 
Invasions of Port Phillip Bay, Victoria . Centre for Research on Introduced 
Pests. Technical Report No. 20. CSIRO Marine Research; Hobart, Australia; 
344 pp. 
Inglis, G J; Floerl, O; Unwi n, M; Ponder-Sutton, A; Seaward, K; Kospartov, M; Bell, 
A; Kluza, D (in prep.). The biosecurity risks associ ated with biofouling on 
international vessels arriving in New Zealand. Summary of the patterns and 
predictors of fouling.  Biosecurity New Zealand Technical Paper. 
Johns, P; Newman, L J; Holleman, J J; Da wson, E W; Sterrer, W; Allison, F R; 
Diggles, B K; Andrews, J R H; Hine, P M; McKenna, P B; Poulin, R (2009) 
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  93  
Phylum Platyhelminthes. In Gordon, D P (ed) New Zealand Inventory of 
Biodiversity. Volume 1 : Kingdom Animalia; Radiata, Lophotrochozoa, 
Deuterostomia. Canterbury University Press; Christchurch, New Zealand; pp 
102-128. 
Lee, K-M (2006) Taxonomy and ecology of predatory marine flatworms 
(Platyhelminthes: Polycladida) in Bo tany Bay, New South Wales, Australia . 
MSc thesis; University of New South Wales, Australia. 
Lee, K-M; Beal, M A; Johnston, E L (2006) a new predatory flatworm 
(Platyhelminthes, Polycladida) from Botany Bay, New South Wales, Australia. 
Journal of Natural History  39(47): 3987-3995. 
Minchin, D; Gollasch, S (2003) Fouli ng and Ships’ Hulls: how Changing 
Circumstances and Spawning Events may Result in the Spread of Exotic Species. Biofouling  19: 111-122. 
Newman, L J; Cannon, L R G; Govan, H (1993) Stylochus ( Imogene ) matatasi  n. sp. 
(Platyhelminthes, Polycladida): pest of  cultured giant clams and pearl oysters 
from the Solomon Islands. Hydrobiologia 257: 185-189. 
O’Connor, W A; Newman, L J (2001) Ha lotolerance of the oyster pradator, Imogine 
mcgrathi , a stylochid flatworm from Po rt Stephens, New South Wales, 
Australia. Hydrobiologia  459: 157-163. 
O’Connor, W A; Newman, L J (2003)  Predation of cultured mussels, Mytilus 
galloprovincialis , by stylochid flatworms , Imogine mcgrathi , from Twofold 
Bay, New South Wales, Australia. Asian Fisheries Science  16: 133-140. 
Rawlinson, K A (2008) Biodiversity of coasta l polyclad flatworm assemblages in the 
Wider Caribbean. Marine Biology  153: 769-778. 
Rivest, B R; Coyer, J; Tyler, S (1999)  The first known invasion of a free-living 
marine flatworm. Biological Invasions  1: 393-394. 
Ruppert E E; Fox R S; Barnes R D (2004) Invertebrate Zoology: A functional 
evolutionary approach  (7th edition). Thompson Lear ning; Victoria, Australia. 
Skerman TM (1960) Note on Stylochus zanzibaricus  Laidlaw, (Turbellaria, 
Polycladida), a suspected predator of barnacles in the Port of Auckland, New 
Zealand. New Zealand Journal of Science 3: 610-614. 
Sluys, R; Faubel, A; Rajagopal, S; Velde,  G van der (2005) A ne w and alien species 
of “oyster leech” (Platyhelminthes, Polycladida, Stylochidae) from the brackish North Sea Canal, The Netherlands. Helgoland Marine Research  59: 310-314. 
Species 2000. Catalogue of Life: Dynamic Checklist . 
http://www.catalogueoflife.org/dynamic- checklist/browse_taxa.php. Accessed 3 
March 2009. 
 
94  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  95 4.10. GASTROPODS AND CHITONS  (CLASS: GASTROPODA AND 
POLYPLA
COPHORA) 
4.10.1.  Hazard Identification 
Gastropoda (gastropods) and Po lyplacophora (chitons) are tw
o classes of non-bivalve 
molluscs. The Gastropoda are a diverse class of mobile snail- or slug-like organisms 
primarily adapted to benthic life. They are divided into the subclasses of 
Prosobranchia, Opistobranchia and Pulmonata.  The prosobranchs are marine, freshwater and terrestrial snails that usually have a coiled shell and operculum
12. By 
contrast, opistobranchs have a shell that is uncoiled, reduced or entirely absent, e.g. 
sea slugs, sea hares and sea butterflies. Subclass Pulmona ta includes a few marine 
intertidal species, such as limpets, as well as  land snails and many freshwater forms.   
 The chitons are small, armoured, herbivorous  molluscs that live attached to hard 
substrata by a broad flat foot . They are dorsoventrally flattened and oval-shaped, with 
eight overlapping calcareous plates (Ruppert  et al. 2004). When immersed, chitons 
graze algae from hard substrata. Many chiton species have a restricted range, returning to the same niche in the substrat a before being emmersed at low tide. The 
chiton makes a watertight seal protecting it against desiccation and dislodgement (Ng 
& Williams 2006).  Reproduction varies among the gastropod s ubclasses. Prosobranchs are gonochoric 
with external fertilisation while the hermaphroditic opisthobranchs and pulmonates copulate and fertilise internal ly (Ruppert et al. 2004).  The majority of gastropods lay 
gelatinous egg masses, some hatching as pl anktonic veliger larvae while others hatch 
juveniles.  Chitons have separate se xes and reproduce via broadcast spawning. 
 Ecologically, the grazing of hard surf aces by chiton and herbivourous gastropods 
influences the structure and diversity of fouling communities (Duffy & Hay 2001, 
Aguilera & Navarrete 2007, Nydam & St achowicz 2007). In addition, predatory 
gastropods may contribute to the shaping of sessile fouling communities (Nydam & 
Stachowicz 2007), while detritivorous gastropods disperse the soft sediments in which 
they feed, incorporating the oxygen that is essential for infaunal organisms. 
 Globally, there have been many recorded in troductions of gastropods, particularly, 
opistobranchs (Boyd 1999, Albayrak & Cagl ar 2006, Ozturk & Can 2006, Daskos & 
Zenetos 2007). The impacts of these inva sions include economic losses and changes 
to the benthic community, for example, lo ss of diversity and alteration of trophic 
webs. Large predatory whelk species ( Rapana venosa, Ocinebrallus inornatus  and 
Urosalpinx cinerea ) threaten wild and commercial biva lve populations in a variety of 
locations (Mann & Harding 2000, Martel et al. 2004, Savini & Occhipinti-Ambrogi 
2006, Faasse & Ligthart 2007). The introductio n of the unusual filter-feeding limpet, 
Crepidula fornicata,  has resulted in a vari ety of environmental impacts (Richard et al. 
2006).  
                                                
 
12 A round ‘lid’ used to block the opening once the organism has retracted into the shell.  
There are two recorded cases of non-indigenous chitons forming established 
populations. Although the “Japanese chiton” Acanthichitona achates  was 
inadvertently introduced to Washington State, USA in 1948 through seed oysters, no impacts have been recorded (Ray 2005, Elder 2009). However, the chiton, Chiton 
glaucus,  introduced to southeaste rn Tasmania prior to 1910, has become “one of the 
most conspicuous and common chiton species and hence must be having some impact 
on the native species” (Hayes et al. 2005).  Given that gastropods and chitons: 
 are known components of marine  biofouling assemblages; 
 are known to have been introduced to new locations; 
 have known impacts on core values; 
it is concluded that gastropod and chiton sp ecies represent a potential hazard via 
vessel biofouling. 
 
4.10.2.  Risk Assessment 
4.10.2.1.  Entry assessment 
Gastropods and chitons are known com
ponents of vessel biofouling assemblages. For 
example, six gastropods and one vermitid sp ecies were recorded in a hull fouling 
survey at the Port of Recife, Brazil (Farra peira et al. 2007).  Two species of gastropod 
have established in New Zealand after being introduced as fouling on oil rigs 
(Cranfield 1999).  In the MAFBNZ commissioned research on biofouling, 
approximately 1% of vessels that entered New Zealand had either gastropods and/or 
chitons recorded (Inglis  et al. in prep.).  
 The gastropods and chitons usually found in  association with vessel biofouling are 
small intertidal/upper subtidal  species with high attachme nt strength and/or close 
association with other hull fouling species (Cranfield et al. 1999, Mann and Harding 
2000). For example, many nudibranchs introduc ed to New Zealand and Port Phillip 
Bay (Australia) have been associated with the hull fouling bryozoans upon which they 
feed (Miller and Willan 1986 In Boyd 1999, Cranfield 1999). Nudibranchs are 
capable of swimming directly onto the hull of  a vessel and it has been suggested that 
egg masses attached to a hull may result in introductions (Kerckhof et al. 2006, Faasse 
& Ligthart 2007). In gene ral, adults of the larg er predatory molluscs and C. fornicata  
occupy infaunal habitats, however, juveniles spe nd the first part of their life in fouling 
communities on hard surfaces. Within the hull environment, seachests may provide the most suitable habitat fo r gastropods and chitons. Depending on their life-history, 
they could either settle independently or be drawn in by the action of the seachest (Mann & Harding 2000).  Gastropods and chitons have a known ability to survive translocation. For example, 
many of the introduced gastropods recorded in  New Zealand are native to either Japan 
or Europe (Cranfield 1999). Gastropods with  shells are able to retreat and plug the 
aperture with mucous or their opercul um, protecting themselves from adverse 
physical conditions. In addition, many of the molluscs found in hull fouling are from coastal habitats and are thus adapted to changes in environmental conditions. For 
example, Rapana venosa  adults can tolerate salinities in the range of 18-28 ppt whilst 
96  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
juveniles show no effects when exposed to salinities as lo w as 10 ppt (Mann & 
Harding 2000).  
 There have been no known introductions of non-indigenous gastropods or chitons in 
New Zealand since the early 1980’s (Kospartov et al. in press).  
 The effect of anti-fouling paint on gastr opods and chitons is un certain. The biocide 
tributyltin was responsible for the near exti nction of certain wh elk species (Faasse & 
Ligthart 2007), however, this biocide ha s been banned internationally since 2008.  
As modern anti-fouling paints require regular current movement to be effective (Floerl et al. 2005), it is possi ble that gastropods and chitons may settle in the niche 
areas of the hull where the water current results in suboptimal hull protection. In 
addition to the physical barrier provided by mucous production, close association with other biofouling is also exp ected to diminish contact between these organisms and 
anti-fouling biocides.  Given that particular ga stropod and chiton species: 
 are known to occur along vessel pathways; 
 are known to be associated with vessel biofouling; 
 are known to survive translocation; 
it is concluded that the likelihood of en try of non-indigenous gastropod and chiton 
species into New Zealand via ve ssel biofouling is non-negligible. 
4.10.2.2.  Establishment assessment 
Suitable habitat for non-indigenous gastropods  and chitons is known to exist in New 
Zealand. The prevalence of hard substrate in ports and m
arinas within New Zealand 
may provide suitable substrate for establishment of gastropods and chitons. In event of the introduction of predatory whelks or C. fornicata , the harbours close to ports and 
marinas commonly encompass estuarine a nd marine infaunal (muddy/sandy) habitats. 
 Gastropods and chitons can ente r and establish through the m ovement of adults and/or 
juveniles directly off the vessel, for ex ample nudibranchs, or through spawning and 
settlement. The dislodgement of closely asso ciated biofouling or individuals may also 
lead to the introduction of non-indigenous ga stropods and chitons to New Zealand. As 
most of the gastropods found on vessel biofoulin g are likely to be broadcast spawners 
(Mann and Harding 2000), for fertilisation to  take place there must be at least two 
individuals in close proximity. Pelagic larv al phases allow for long distance dispersal 
and establishment in other locations (G ilman 2006, Viard et al. 2006) by increasing 
the chances of encountering suitable ha bitat. Some snails and many nudibranchs 
associated with ephemeral biofouling have  short life-cycles and may mature in a 
matter of days or weeks (Willan 2003). The shorter life-cycle may increase the chances of establishment in New Zealand wa ters as species are more likely to be 
reproductively active following entry into  New Zealand. The sex and reproductive 
output of some gastropod species is depende nt on the chemicals released by other 
conspecifics, for example, the establishment of C. fornicata is reliant on the presence 
of other juveniles or adults in  the near vicinity (ISSG 2005). 
 The rates of non-indigenous sp ecies establishment can ofte n be correlated with the 
abundance of the organisms at the source lo cation. For example, Miller et al. (2007) 
investigated the successful and failed mollu scan invaders in San Francisco Bay in 
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  97  
association with oyster transf ers from the east coast of th e USA. They found that the 
relative abundance of mollusc s in the source region was th e strongest predictor of 
their invasion success, thus emphasising the importance of propagule pressure to 
invasion success (Mil ler et al. 2007). 
 The most recent known establishment of a gastropod or chiton in New Zealand was the nudibranch Janolus hyalinus prior to 1981 (Kospartov et al. in press). There have 
been no recorded establishments of chit ons in New Zealand (Kospartov et al. in 
press).   Given that particular ga stropod and chiton species: 
 are known to inhabit similar climates/env ironments to those that exist in New 
Zealand; 
 are capable of forming establishi ng populations in New Zealand; 
 have been previously recorded as established in New Zealand (gastropods 
only); 
it is concluded that the likelihood of es tablishment of non-indigenous gastropod and 
chiton species in New Zealand via vessel biofouling is non-negligible. 
4.10.2.3.  Consequence assessment 
The consequences of herbivorous gastropods or chitons in New Z ealand are likely to 
be inte
rspecific competition with othe r grazers. A number of common grazing 
gastropods and chitons of various sizes ar e present in New Zealand coastal waters, 
including C. glaucus (Morton 2004). This chiton is know n to exhibit intraspecific 
competition and interspecific competition with the omnivorous seastar Patiriella 
regularis  in intertidal cobble fields (Palmer & Phillips 2009). It is uncertain whether 
non-indigenous introductions will compet e, be excluded, co-exist, or find a new 
niche. 
 Ecological consequences of predatory gast ropods in New Zealand are likely to be 
interspecific competition with other predators, and possible negative effects on indigenous fauna. The establishment of pr edatory gastropods may negatively impact 
on bivalve aquaculture.     Oyster drills are gastropods of the fam ily Muricidae that prey on oysters, other 
bivalves, including mussels (Hanks 1957), and barnacles (Faasse & Ligthart 2007). 
Among pest species, oyster drills are widely considered to have the highest impacts to 
world oyster production. These organisms can prey on oysters of any size and impacts can range from nil to over 90% predati on depending on local conditions (Butler 
1985). Specifically, the American oyster drill ( Urosalpinx cinerea ) was recorded as 
consuming more than half the oyster spat in certain European  estuaries (Cole 1942 In 
Hulme et al. 2008). In 2003, the global value of oyster (Crassostrea gigas ) production 
was estimated at US$3.69 billion (FAO 2006). Despite enormous research efforts, there appears that little can be done to mitigate the effects of oyster drills once 
established, the only mitigation measures ar e manually scrubbing snail eggs off the 
oysters, leaving areas fallow for periods of time, or growing in less saline areas 
(National Sea Grant College Program 2003).  Given that the prey of oyster drills can include both oysters and mussels, and the known impacts to commercial stocks in othe r jurisdictions, the pot ential consequences 
98  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
to New Zealand’s aquaculture industry could be wide-ranging. In 2008, mussels 
(Perna canaliculus) and oysters ( C. gigas ) comprised of 77% (NZ$204.3 million) and 
6% of the value (NZ$16.9 million) of New Zealand’s aquaculture exports, 
respectively (Aquaculture New Zealand 2009). Given the value of bivalve aquaculture 
and commercial fishing to New Zealand, the possible establishment of specific 
gastropod species via hull biofouli ng is a non-negligible concern.  
 
The slipper limpet C. fornicata  is a filter feeding ga stropod that forms large 
aggregations and may occur at densities exceeding 1700 m-2 (DAISIE 2011). These 
structures have been reported to change  local sedimentation dynamics, impacting the 
benthic community.  In additi on, their filter feedi ng behaviour has been implicated in 
recruitment failures for commercial oyster species and declines of juvenile sole (commercially harvested Solea solea ) in nursery habitats (Le Pape et al. 2004, 
Richard et al 2006, Beninger et al 2007, Deco ttignies et al 2007). Further, slipper 
limpets require removal from oyster and s callops before marketing and as such 
increase commercial costs (Minchin 2008). Introductions  of this species have 
occurred through shellfish stock movement and vessel biofouling (DAISIE 2011). 
4.10.3.  Risk Estimation 
Given that particular gastropod species: 
 are known com
ponents of hull fouling assemblages;  
 are known to have been introduced to new locations; 
 are capable of forming establishi ng populations in New Zealand; 
 have well known invasive behaviours; 
 have had significant im pacts on core values; 
 are likely to have adverse effects on core  values, such as, bivalve aquaculture; 
it is concluded that the import of non-i ndigenous gastropod species via biofouling 
poses a non-negligible risk to Ne w Zealand’s marine core values. 
 
Given that particular chiton species: 
 are not common components of vess el hull biofouling assemblages; 
 have not been recently recorded as introduced to New Zealand; 
 are not recorded as being responsib le for impacts to core values;  
it is concluded that the import of non-indi genous chiton species via biofouling poses a 
negligible risk to New Zeal and’s marine core values. 
 
4.10.4.  Risk Management 
A generic approach is suitable to m
anagi ng the risk associated with gastropods 
(Section 3). In addition, the mobility of gastropods and the dislodgement of their egg 
masses present further management difficulties,  as a result, systems that contain all 
organisms in-water prior to treatment ma y be required to reduce the risk of pre-
treatment escape. 4.10.5.  References 
Aguilera, M A; Navarrete, S A (2007) Effects of Chiton granosus  (Frem 
bly, 1827) 
and other molluscan grazers on algal succession in wave exposed mid-intertidal 
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  99  
rocky shores of central Chile. Journal of Experimental Marine Biology and 
Ecology 349: 84-98. 
Albayrak, S; Caglar, S (2 006) On the presence of Siphonia belcheri Hnaley, 1858 
[Gastropoda: Siphonariidae] and Septifer bilocularis  (Linnaeus, 1758) 
[Bivalvia: Mytilidae] in the Iskenderun Bay (SE Turkey). Aquatic Invasions  
1(4): 292-294. 
Aquaculture New Zealand (2009) New Zealand Aquaculture farm facts (2nd edition). 
http://www.aquaculture.org.nz/aquacultu re-in-nz/industry-overview/overview/ . 
Accessed 20 January 2010. 
Beninger, P G; Decottignies, P; Guihen euf, F; Barille, L; Rince, Y (2007) 
Comparison of particle processing by  two introduced suspension feeders: 
selection in Crepidula fornicata and Crassostrea gigas .  Marine Ecology 
Progress Series  334:165-177. 
Boyd, S (1999) The Introduced Mollusca of Port Phillip Bay.  In: Hewitt C L; 
Campbell M L; Thresher R E; Martin R B (Eds.) Marine Biological Invasions of 
Port Phillip Bay , Victoria .  Centre for Research on Introduced Marine Pests. 
Technical Report No. 20. CSIRO Marine Research; Hobart, Australia; pp.129-
149. 
Butler, P A (1985) Synoptic review of the literature on the Southern Oyster Drill 
Thais haemastoma floridana . NOAA Technical Report NMFS 35. 9pp. 
Cranfield, H J; Gordon, D P; Willan, R C; Mars hall, B A; Battershill, C N; Francis, M 
P; Nelson, W A; Glasby, C J; Read, G B (1998) Adventive marine species in 
New Zealand. NIWA Technical Report 34. 48 pp. 
DAISIE European Invasive Alien Species Gateway, 2011. Crepidula fornicata. 
Available from:  
http://www.europealiens.org/spec iesFactsheet.do ? speciesId=50159#  Accessed 5 
January 2011. 
Daskos, A; Zenetos, A (2007) Additions to the knowledge of alie n Opistobranchia of 
Greece. Aquatic Invasions  2(3): 258-260. 
Decottignies, P; Beninger, P G; Rince, Y; R obins, R J; Riera, P (2007) Exploitation of 
natural food sources by two sympatri c, invasive suspension feeder: Crassostrea 
gigas  and Crepidula fornicata.  Marine Ecology Progress Series  334: 179-192. 
Duffy, J E; Hay, M E (2001) The ecology a nd evolution of marine consumer-prey 
interactions. In Bertness, M D; Hay, M E; Gaines, S D (eds) Marine Community 
Ecology. Sinaur Associates; Sunderland, United States of America; pp 131-157. 
Elder, N (2009) Acanthochitona achates. USGS Non-indigenous Aquatic Species 
Database. http://nas.er.usgs.gov/queries/FactSheet.asp?speciesID=1190. Accessed 24 September 2009 
Faasse, M; Ligthart, M (2007) The American oyster drill, Urosalpinx cinerea  (Say, 
1822) introduced to the Netherlands – increased risks after ban on TBT?  
Aquatic Invasions  2: 402-406. 
FAO (2006) Crassostrea gigas  fact sheet: Cultured Aquatic Species Information 
Program. http://www.fao.org/fishery/culture dspecies/Crassostrea_gigas/en
. 
Accessed 3 February 2010. 
Farrapeira, C M R; Melo, A V de O M de; Ba rbosa, D F; Silva, K M E da (2007) Ship 
hull fouling in the Port of Recife, Pernambuco.  Brazilian Journal of 
Oceanography  55: 207-221. 
Gilman, S E (2006) The northern geographic range limit of the in tertidal limpet 
Collisella scabra : a test of performance, recruitment, and temperature 
hypotheses. Ecogeography  29: 709-720. 
100  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
Hanks, J E (1957) The feeding rates of the common oyster drill, Urosalpinx cinerea  
(Say), at controlled water temperatures. Biological Bulletin  112:330-335. 
Hulme, P E; Roy, D B; Cunha, T; Larsson, T-B (2008) A pan-Eur opean inventory of 
alien species: rationale, implementa tion and implications for managing 
biological invasions. In DAISIE (eds.) The Handbook of European Alien 
Species . Springer, Dordrecht. 
Inglis, G J; Floerl, O; Unwi n, M; Ponder-Sutton, A; Seaward, K; Kospartov, M; Bell, 
A; Kluza, D (in prep.). The biosecurity risks associ ated with biofouling on 
international vessels arriving in New Zealand. Summary of the patterns and 
predictors of fouling.  Biosecurity New Zealand Technical Paper. 
Kospartov K; Inglis G; S eaward K; Brink A van den;  D’Archino R; Ahyong S (in 
press) Non-indigenous and cryptogenic marine species in New Zealand . Current 
state of knowledge – Interim report pr epared for MAFBNZ Post Border 
Directorate. MAF Biosecurity Ne w Zealand. Wellington, New Zealand. 
Le Pape, O: Guerault, D; Desaunay, Y ( 2004) Effect of an invasive mollusc, 
American slipper limpet Crepidula fornicata, on habitat suitability for juvenile 
common sole Solea solea in the Bay of Biscay. Marine Ecology Progress Series  
277: 107–115. 
Mann, R; Harding, J M (2000) Invasion of th e North American Atlantic coast by a 
large predatory Asian mollusc. Biological Invasions  2: 7-22. 
Martel, C; Guarini, J M; Bl anchard, G; Sauriau, P G; Trichet, C; Robert, S; Garcia-
Meunier (2004) Invasion by the marine gastropod Ocinebrellus inornatus  in 
France. III. Comparison of biological traits with the resident species Ocenebra 
erinacea .  Marine Biology 146: 93-102. 
Miller, A W; Ruiz, G M; Minton, M S;  Ambrose, R F (2007) Differentiating 
successful and failed molluscan invaders in estuarine ecosystems.  Marine 
Ecology Progress Series  332: 41-51. 
Morton, J (2004) Seashore ecology of New Zealand and the Pacific . David Bateman, 
Auckland, New Zealand. 
National Sea Grant College Program (2003) Th e Gulf Oyster Industry: seizing a better 
future. Communictions office, Louisian a Sea grant College Program; Louisiana 
State University, Louisiana; USA. 17pp. 
Ng, J S S; Williams, G A (2006) Intraspecific variation in foraging behaviour: 
influence of shore height on temporal organization of activity in the chiton 
Acanthopleura japonica .  Marine Ecology Progress Series  321:183-192. 
Nydam, M; Stachowicz, J J (2007) Pr edator effects on fouling community 
development. Marine Ecology Progress Series  337: 93-101. 
Ozturk, B; Can, A (2006) Indo- Pacific gastropod species in the Levantine and Aegean 
Seas. Aquatic Invasions  1(3): 124-129. 
Palmer, S; Phillips, N (2009) The ecological role of an omnivorous seastar ( Patiriella 
regularis ) in intertidal cobble fi elds. In O’Shea, S; Cost ello, M J; O’Connell, K 
(eds) Programme and Abstracts for A nnual Conference in Auckland 2-4 
September 2009 . New Zealand Marine Scienc es Society
; Accessed at 
http://nzmss.rsnz.org ; 1-126 pp.  
Ray, G L (2005) Invasive marine and estuar ine animals of the Pacific northwest and 
Alaska. U.S. Army Engineer Research  and Development Center; Vicksburg, 
U.S.A.; 19 pp. 
Richard, J; Huet, M; Thouzea u, G; Paulet, Y-M (2006) Re production of the invasive 
slipper limpet, Crepidula fornicata, in the Bay of Brest, France. Marine Biology  
149: 789-801. 
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  101  
102  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis Ruppert, E E; Fox, R S; Barnes, R D (2004) Invertebrate Zoology: A functional 
evolutionary approach  (7th edition). Thompson Lear ning; Victoria, Australia. 
Savini, D; Occhipinti-Ambrogi, A (2006) C onsumption rates and prey preference of 
the invasive gastropod Rapana venosa in the Northern Adriatic Sea.  Helgoland 
Marine Research  60: 153-159. 
Viard, F; Ellien, C; Dupont, L (2006) Disp ersal ability and invasion success of 
Crepidula fornicata in a single gulf: in sights from genetic markers and larval-
dispersal model. Helgoland Marine Research  60: 144-152. 
Willan, R (2003) Nudibranchs . In Andrew, N; Francis, M (eds) The Living Reef – the 
ecology of New Zealand’s rocky reefs. Craig Potton Publishing; Nelson, New 
Zealand.  
4.11. HYDROIDS (CLASS: HYDROZOA) 
4.11.1.  Hazard Identification 
Hydrozoa are a diverse class of predator y anim
als belonging to the phylum Cnidaria. 
This class consists of approximately 3,000 species, the majority of which occupy 
marine habitats (Ruppert et al. 2004). Hydr ozoans have two alternative body plans: 
the polyp and the medusa. A typical polyp is a sedentary organism that has the mouth 
uppermost which is surrounded by tentacles laden with stinging cells. The polyp body 
plan is represented by the sea anemone. A typical medusa (upside-down polyp) has 
the mouth on the underside of a bell that is surrounded by a trail of tentacles. These 
species swim by pulsing of the bell. Th e medusa body plan is represented by the 
jellyfish. Polyps often reproduce by budding off other polyps, however, in some 
species these buds remain attached to fo rm colonies. Coloni al species include 
seaweed-like hydroids, pelagic Portuguese man-of-war, and r eef building fire and rose 
corals. The term hydroid refers to sessile benthic colonies, while  colonial pelagic 
species are called siphonophores (Ruppert et al. 2004, Cairns et al. 2009).  
 The life-cycle of hydrozoa involves a benthic polyp phase and a pelagic medusa phase. Hydroid colonies form from an  individual polyp by budding asexually. A 
mature colony consists of both feeding polyps , with tentacles and stinging cells, and 
reproductive polyps without tentacles, which regularly bud off medusae (Ruppert et al. 2004). The medusae are the dispersal and se xual stage of the life-cycle. Cells in 
mature medusae undergo meiosis and give ri se to either sperm or eggs, which are 
released into the surrounding water to be fertilised. Af ter development, the larva 
settles onto a suitable  substrate to give rise to a new colony (Ruppert et al. 2004).  
 Ecologically, hydrozoans are successful colonists and can have high abundances 
under certain environmental conditions (M a & Purcell 2005). Hydroids constitute an 
important component of the fouling comm unity, growing rapidly and profusely on 
both natural and anthropogeni c hard substrata. As mi cro-carnivores it has been 
suggested that hydroids may play an importa nt role in the ener gy transfer between 
benthos and plankton (Ma & Purcell 2005). The hydromedusae produced by the 
polyps have also been shown to impact  zooplankton and ichthyoplankton populations 
(Larson 1987, Purcell & Grover 1990). Due to their effect s on larval settlement and 
community structure, the presence of hydroi ds on vessel hulls is a concern for both 
aquaculture (Mortensen et al. 2000) and wildlife herita ge areas (Wyatt et al. 2005, 
Lewis et al. 2006). Well known intro duced hydrozoan species include Obelia 
dichotoma, Moerisia lyon si, Maeotias marginata, Blackfordia virginica (Cohen & 
Carlton 1995, Vainola & Oulasvirta 2001, Ma & Purcell 2005, Bardi & Marques 
2009).  
Given that hydroids: 
 are known components of marine  biofouling assemblages;  
 are known to have been introduced to new locations; 
 have known impacts on core values; 
it is concluded that hydroid species represen t a potential hazard vi a vessel biofouling. 
 
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  103  
4.11.2.  Risk Assessment 
4.11.2.1.  Entry assessment 
Hydroids are common fouling com
ponents in ports, on wharf piles, and other 
anthropogenic surfaces (Lewis 2004, Cairns et al. 2009). Therefore, hydroids are 
likely to encounter the hulls of international vessels destined for New Zealand. 
 Hydroids are common components of hull f ouling. For example, Woods Hole (1952) 
listed more species of hydroids than any other macrofaunal hull fouling group.  Hydroids can be found on hulls with or wit hout other biofouling and are tolerant to 
high concentrations of coppe r ions, the active ingredient of the majority of anti-
fouling paints (Pyefinch & Downing 1949 In Lewis 2004, Railkin 2004). Hydroids 
have traditionally been f ound in greater numbers on slower moving vessels and on 
hull locations that experience slower fl ow velocities (Skerman 1960, Railkin 2004). 
For example, in the MAFBNZ commissioned research hydrozoan fouling was present 
on 8% of sampled vessels, 78% of which were recreational vessels, a slow moving vessel class.  Hydroids are known to survive translocati on. The “contamination” of Pacific coast 
harbours by ship-introduced hydroids wa s recorded as early as 1915 (Cohen & 
Carlton 1995). To be transported by vessels , species must occur in estuarine and 
coastal habitats and therefore be tolerant  of a range of salinities and high water 
turbidity (Watson 1999). Cryptic and sma ll hydroids are more likely to survive 
voyages as they are likely to be sheltered by larger fouling organisms or be present in 
niche areas (Watson 1999). The la rval stages of some species can also be cryptic and 
it has been suggested that la rger species could be transl ocated via this lifestage 
(Watson 1999). Until recently, coincident wi th the increased interest in marine 
biosecurity, there were few data regarding the transport of hydroids via shipping 
(Watson 1999). 
 Taxonomic identification of hydrozoans is pr oblematic. Difficulties  in identification 
may be due to species similarities, high colony variability or poor specimen quality, 
for example, distinguishing features may be  missing, particularly towards the end of 
the reproductive season and as predators and currents take their effect (Boyd et al. 
2002). As a result there has been a history of misidentifications, for instance in San Francisco Bay, New England, and Europe (Cohen & Carlton 1995, B oyd et al. 2002).  
In the MAFBNZ commissioned research on biofouling, 22 of 58 hydrozoan samples 
were classified as indetermin ate (Inglis et al. in prep.). 
 Given that particular hydroid species: 
 are known to occur along vessel pathways; 
 are known to be associated with vessel biofouling; 
 are known to survive translocation; 
it is concluded that the likelihood of en try of non-indigenous hydroid species into 
New Zealand via vessel biof ouling is non-negligible. 
 
4.11.2.2.  Establishment assessment 
Hydroids are general opportunist  foulers of hard substrates and, as such, are likely to 
find suitable and available habitats in New Zealand. W
atson ( 1999) noted that the 
104  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
majority of fouling species would succeed in the high flow environments provided by 
the jetty pilings of port and harbours. The temperature range  of the receiving 
environment is likely to be an important influence of the establishment of hydroid 
populations, particularly for seasonal species  that are adapted to narrow bands of 
summer or winter temperatur es (Watson 1999). Given their wide tolerances, salinity 
and turbidity are unlikely to be major influences.  
 Hydroids can establish in the New Zeal and environment through fragmentation of 
colonies (asexual reproduction) or the re lease of medusae (sexual reproduction). In 
ports or harbours, fertile col onies or fragments may be rel eased into the water column 
by abrasion during berth or during hull cleaning operations . The release of medusae 
may occur if fertile colonies are component s of the biofouling assemblage. As smaller 
hydroid species grow and reach maturity fast er than larger species, they are more 
likely to pass through a life-cycle and re lease propagules in New Zealand waters 
(Watson 1999).  Twenty six species of non-indigenous hydroi d are recorded as established in New 
Zealand (Kospartov et al. in press). Five of these species have  established since 2000. 
Non-indigenous hydroids have been r ecorded throughout New Zealand from the 
Auckland Islands to the Kermadec Isla nds (Kospartov et al. in press). 
 Given that particular hydroid species: 
 are known to inhabit similar climates/env ironments to those that exist in New 
Zealand; 
 are capable of forming establishi ng populations in New Zealand; 
 have been recently recorded as  established in New Zealand; 
it is concluded that the likelihood of esta blishment of non-indige nous hydroid species 
in New Zealand via vessel biofouling is non-negligible.  
4.11.2.3.  Consequence assessment 
The consequences of the establishm
ent of non-indigenous hydroids to New Zealand is 
likely to range from nuisance fouling to the structuring of benthic fouling 
communities. The fouling of vessel hulls and general water intake pipes may lead to increased maintenance, however, this shoul d be resolved by a regular schedule and 
should not result in signif icant economic impacts.  
 Ecologically, hydroids can dominate parts of the sea floor; as micro-carnivores they 
may have unforeseen effects on New Zeala nd marine assemblages. Hydrozoans feed 
on a range of eggs, larvae, and small fauna present in the water column, transferring 
energy from the pelagic environment to th e benthos (Gili et al. 1996). Moreover, their 
predation of eggs and larvae in the water column can structure the benthic fouling 
community. For example, Standing (1976) f ound that the predation of the settling 
barnacle cyprids ( Balanus crenatus ) by Obelia dichotoma  in Bodega Bay, California, 
USA, allowed the settlement of the otherwise excluded solitary ascidian ( Ascidia 
ceratodes ). In addition, Wyatt et al. (2005) expressed concern over introduced 
hydroids in the Shark Bay World Heritage S ite, Australia, because of their effects on 
larval settlement.   
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  105  
Consequences of non-indigenous species of hyd roids are likely to include fouling of 
anthropogenic structures in coastal marine se ttings. Hydroids are a substantial part of 
aquaculture biofouling assemblages. The imp act of fouling of aquaculture operations 
may include the inhibition of bivalve gr owth, competition for resources and the 
consumption of settling larvae (Sandifer et al. 1974, Claereboudt et al. 1994). The 
costs of biofouling to the aquaculture i ndustry can be significant. For example, 
biofouling can reduce bivalve growth by ove r 40% and the cleani ng of oyster cultures 
is estimated account for 20% of their market  value. The annual cost of fouling on 
cultured mussels in Scotland is estima ted to be up to 750,000 Euro (Campbell & 
Kelly 2002). Hydroid colonies can recover within a week of cleaning, with some 
colonies having increased thickness (Guenther et al. in press).  
 
As hydroids provide habitat a nd chemical cues for a wide  range of other organisms 
such as amphipods and polychaetes, they may facilitate the settlement of other non-
indigenous species in port/marina and aquacu lture environments which may result in 
further consequences to core values. 
4.11.3.  Risk Estimation 
Given that particular hydrozoan species: 
 have been recorded entering New Zealand via hull fouling; 
 are unlikely to be correctly iden tified upon entry into New Zealand; 
 m
ay facilitate the entry of a variety of other non-indigenous organisms; 
 are capable of forming establishi ng populations in New Zealand; 
 have known impacts to economic values, such as, aquaculture production; 
 have the potential to impact benthic community structures with far reaching 
indirect effects; 
it is concluded that the import of non-i ndigenous hydrozoan species via biofouling 
poses a non-negligible risk to Ne w Zealand’s marine core values. 
4.11.4.  Risk Management 
Due to their sm
all and cryp tic nature and the high lik elihood of non-identification, a 
generic approach to the removal of bi ofouling would be th e appropriate risk 
management option of hydrozoans (Secti on 3). Specific considerations for hull 
fouling hydrozoans should consider appr opriate local containment due to the 
likelihood of release of associated non-i ndigenous organisms into the receiving 
environment. 
 
4.11.5.  References 
Bardi, J; Marques, A C (2009) The invasive hydrom
edusae Blackfordia virginica  
Mayer, 1910 (Cnidaria: Blackfordiidae) in southern Bazil, with comments on 
taxonomy and distribution of the genus Blackfordia . Zootaxa  2198: 41-50. 
Boyd, M J; Mulligan, T J; Shaughnessy, F J (2002) Non-Indigenous of Humboldt 
Bay, California. A Report to the Califor nia Department of Fish and Game. 
February 28, 2002. 
Cairns, S D; Gershwin, L-A; Brook, F J;  Pugh, P; Dawson, E W; Ocana, O V; 
Vervoort, W; Williams, G; Watson, J E; Opresko, D M; Schuchert, P; Hine, M; 
106  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
Gordon, D P; Campbell, H J; Wright, A J; Sanchez, J A; Fautin, D G (2009) 
Phylum Cnidaria. In Gordon, D P (ed) New Zealand Inventory of Biodiversity. 
Volume 1 : Kingdom Animalia; Radiata, L ophotrochozoa, Deuterostomia. 
Canterbury University Press; Ch ristchurch, New Zealand; pp 59-101. 
Campbell, D A; Kelly, M S (2002) Settlement of Pomatoceros triqueter (L.) in two 
Scottish lochs, and factors determing its abundance on mussels grown in 
suspended culture. Journal of Shelfish Research  21: 519-527. 
Claereboudt, M R; Bureau, D; Cote, J; Himmelman, J H (1994) Fouling development 
and its effect on the growth of  juvenile giant scallops ( Placopecten 
magellanicus) in suspended culture. Aquaculture  121(4): 327-342. 
Cohen, A N, Carlton, J T (1995)  Non-indigenous Aquatic Spec ies in a United States 
Estuary: A Case Study of the Biological Invasions of the San Francisco Bay and Delta. A Report for the United States Fi sh and Wildlife Service, Washington 
D.C. and The National Sea Grant Pr ogram, Connecticut Sea Grant (NOAA 
Grant Number NA36RG0467). 
Gili, J-M; Hughes, R G; Alva, V (1996) A quantitative study of the feeding of the 
hydroid Tubularia larynx  Ellis and Solander, 1786. Scientia Marina  60(1): 43-
54. 
Guenther, J; Misimi, E; Sunde, L M (in press) The development of biofouling, 
particularly the hydroid Ectopluera larynx , on commercial salmon cage nets in 
Mid-Norway. Aquaculture. 
Inglis, G J; Floerl, O; Unwi n, M; Ponder-Sutton, A; Seaward, K; Kospartov, M; Bell, 
A; Kluza, D (in prep.). The biosecurity risks associ ated with biofouling on 
international vessels arriving in New Zealand. Summary of the patterns and 
predictors of fouling.  Biosecurity New Zealand Technical Paper. 
Kospartov K; Inglis G; S eaward K; Brink A van den;  D’Archino R; Ahyong S (in 
press) Non-indigenous and cryptogenic marine species in New Zealand . Current 
state of knowledge – Interim report pr epared for MAFBNZ Post Border 
Directorate. MAF Biosecurity Ne w Zealand. Wellington, New Zealand. 
Larson, R J (1987) Trophic ecology of plankt onic  gelatinous predators in Saanich 
Inlet, British Columbia: diet and prey selection. Journal of Plankton Research  
9: 811-820. 
Lewis, J (2004) Hull fouling as a vector for the translocation of marine organisms: 
Report 1: Hull fouling research . Ballast Water Research Series Report No. 14. 
Dept of Agriculture, Fisheries and Forestry; Canberra, Australia; 129 pp. 
Lewis, P N; Bergstrom, D M; Whinam J (2006) Barging in: a temperate marine 
community travels to the sub-Antarctic. Biological Invasions 8: 787-795. 
Ma, X; Purcell, J E (2005) Effects of temperature, salin ity, and predators on mortality 
of and colonization by the invasive hydrozoan Moerisia lyonsi . Marine Biology 
147: 215-224. 
Mortensen, S; Van Der Meeren, T; Fosshage n, A; Hernar, I; Harkestad, I; Torkildsen, 
L; Bergh, O (2000) Mortality of scallop sp at in cultivation, infested with tube 
dwelling bristle worms, Polydora  sp.
  Aquaculture International 8: 267-271. 
Purcell, J E; Grover, J J (1990) Predation a nd food limitation as causes of mortality in 
larval herring at a spawning ground in British Columbia. Marine Ecology 
Progress Series  59: 55-61. 
Railkin, A I (2004) Marine Biofouling: Colonization Processes and Defences. CRC 
Press; 303 pp. 
Ruppert, E E; Fox, R S; Barnes, R D (2004) Invertebrate Zoology: A functional 
evolutionary approach  (7th edition). Thompson Lear ning; Victoria, Australia. 
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  107  
108  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis Sandifer, P A; Smith, P I J; Calder, D R ( 1974) Hydrozoans as pe sts in closed-system 
culture of larval decapod crustaceans. Aquaculture 4: 55-59.  
Skerman, T M (1960) Ship-fouling in New Zealand waters: a survey of marine 
fouling organisms from vessels of th e coastal and overseas trade routes. New 
Zealand Journal of Science  3: 620-648. 
Standing, J D (1976) Fouling community structure: effects of the hydroid, Obelia 
dichotoma , on larval recruitment. In Mackie, G O (ed) Coelenterate ecology 
and behaviour. The Third International Symposium on Coelenterate Biology 
held at University of Victoria, Victoria, British Columbia May 10-13 ; Plenum 
Press, New York, USA; pp 155-164. 
Vainola, R; Oulasvirta, P (2001) The first record of Maeotias marginata  (Cnidaria, 
Hydrozoa) from the Baltic Sea: a Pontocaspian invader. Sarsia 86: 401-404. 
Watson, J E (1999) Review of Hydroids Introduced to Victorian Waters. pp. 88-106. 
In Hewitt, C L; Campbell, M L; Thresher, R E; Martin, R B (Eds) Marine 
Biological Invasions of Port Phillip Ba y, Victoria. Centre for Research on 
Introduced Marine Pests. Technical Report No. 20. CSIRO Marine Research; 
Hobart, Australia. 344 pp. 
Woods Hole Oceanographic Institution (1952) Marine Fouling and its Prevention . 
United States Naval Institute; Anna polis, United States of America. 
Wyatt, A S J; Hewitt, C L; Walker, D I; Wa rd, T J (2005) Marine introductions in the 
Shark Bay World Heritage Property,  Western Australia: a preliminary 
assessment, Diversity and Distributions 11(1): 33-44. 
   
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  109 4.12. MACROALGAE 
4.12.1.  Hazard Identification 
Macroa
lgae are attached macroscopic algae rangi ng in size from filaments a few millimetres 
long (e.g. Ulva  spp. Heesch et al. 2007) to large pl ants greater than 45 metres long (e.g. 
Macrocystis pyrifera  Schiel 2003).  While some speci es are of cryptic habit many are 
significant habitat formers, particularly al ong rocky coastlines (S chiel 1990, Schiel & 
Hickford 2001).  Macroalgae can be generally classified as red, brown or green which 
correspond to the phylogenetic groupings of Rhodophyta, Ochrophyta and Chlorophyta, 
respectively (Adams 1994, Algaebase 2008).    Brown kelps, such as Ecklonia radiata and Durvillaea antarctica , are recognised as 
ecosystem engineers that provide  “vertical structure, habitat, nurseries and food” (Schiel 
2003), regulate community composition and are of ten conspicuous as low intertidal and/or 
shallow subtidal forests (Taylor & Schiel 2005 ).  Although the green and red macroalgae are 
generally smaller in size and less conspicuous  than brown algae they remain important 
components of both intertidal and subtidal ecosy stems with some forming specific biogenic 
habitats, such as rhodolith beds.  Green a nd red macroalgae are morphologically diverse 
ranging from calcified, encrusti ng coralline species, through f ilamentous, foliose bushes to 
larger bladed species.  Of the species known to be introduced into New Zealand, most are 
small red or green species (Nelson & Parsons  2003) although significan t exceptions include 
Codium fragile  ssp. fragile , Undaria pinnatifida  and Grateloupia turuturu . 
 A global review of macroalgal species intr oductions provides a list of 277 species with 
representatives from Rhodophyta, Ochrophyta and Chlorophyta (Williams & Smith 2007).  
Arguably the most prominent of thes e species are the Japanese kelp Undaria pinnatifida and 
the aquarium weed Caulerpa taxifolia  (Lowe et al. 2004).  However, there are other species 
reported to have significant invasive impacts, such as, Caulerpa racemosa varieties, 
Sargassum muticum, Grateloupia turuturu  and Codium fragile  ssp. tomentosoides (Scheibling 
& Gagnon 2006, Strong et al. 2006, Mathieson et al. 2008, Piazzi & Balata 2009).    Of the non-indigenous macroalgal sp ecies established in New Zealand U. pinnatifida  is 
arguably the most conspicuous and has had th e greatest recorded im pact, substantially 
modifying subtidal habitat (Battershill et al . 1998, Russell et al. 2008).  Ribera Siguan (2003) 
found most macroalgal translocations could be attributed to vessels and aquaculture. 
Similarly, Williams and Smith (2007) determined that hull fouling was the dominant vector of macroalgal introductions. These factors comb ined account for between 79% (Ribera Siguan 
2003) and 85% (Williams & Smith 2007) of all macr oalgal translocations.  The morphologies 
of macroalgal species associated with hull f ouling are characterised as being small with 
flexible, filamentous vegetative structur es and high growth rate (Ribera Siguan 2003, 
Williams & Smith 2007).  Callow (1986) noted that the dominant macroalgal foulers of 
artificial substrates are of the species Enteromorpha
13 and Ectocarpus, with both taxa having 
forms resistant to anti-fouling biocides. However, larger macroalgae, such as U. pinnatifida , 
have also been associated with hull fouling (Hay 1990).  These larger species may be more 
likely to survive on slower moving vessels, such as  yachts and barges, or are translocated as 
other life stages which can survive on hulls , such as has been demonstrated for the 
microscopic stage of U. pinnatifida  (Coutts 1999).   
 
                                                 
13 Enteromorpha is now considered a synonym of Ulva (Hayden et al. 2003).  
In the MAFBNZ commissioned research on biofou ling, the majority of the identified algal 
species on recreational vessels were small and filamentous w ith only one species not known 
to be established in New Zealand (Inglis et al. in prep.).  However, many of the samples were 
unable to be identified to species level and as a result it was not possible to determine if the 
larger macroalgae were non-indigenous to New Zealand.  The identification of algal samples 
is difficult due to complex taxonomy and the requirement of ephemeral features (e.g. 
reproductive structures) to make conclusive identifications.   
 Given that macroalgae: 
 are known components of marine  biofouling assemblages;  
 are known to have been introduced to new locations; 
 have known impacts on core values; 
it is concluded that macroalgal species repr esent a potential hazard  via vessel biofouling. 
 
4.12.2.  Risk Assessment 
4.12.2.1.  Entry assessment 
Macroalgae are well known hull fouling orga nism
s (Stubbings 1947, Callow 1986a, Mineur et 
al. 2007).  Ribera Siguan (2003) and Williams & Smith (2007) have identified hull fouling as 
the most significant vector for the transloc ation of macroalgal species.  Macroalgae 
establishment on a vessel hull would require a re latively static hull surface, a compromised, 
ineffective or absent fouling control system and a supply of propagules.    Macroalgal spores can settle and attach to the biofilm on a hull.  The initial attachment 
strength is low whereby spor es can easily be dislodged, how ever, once differentiation has 
taken place rhizoid cells provide a considerab ly stronger attachment (Railkin 2004).  The 
spores of some common fouling species can fo rm a strong surface attachment.  For example, 
Ulva  sp. have been found to have propagules that can resist flow forces  equivalent to 40 knots 
within 4 hours of settleme nt (Findlay et al. 2002). 
 Anti-fouling paints are generally effective in  preventing macroalgal settlement, however, 
some species of macroalgae have become tole rant of anti-fouling coatings (Callow 1986b, 
Mineur et al. 2008).  Vessels will have areas of the hull where the an ti-fouling system has 
been compromised (e.g. along the boot top a nd where the hull rubs against pontoons when 
moored) or absent (e.g. sacrific ial anodes and the prope ller).  The majority of macroalgal 
species require an appropriately lit environmen t which restricts their potential zone of 
recruitment to conspicuous areas of the hull, such as just below the waterline.  As a consequence, vessel maintenance for aesthetic and/or performance enhancement will mitigate algal accumulation for most vessels.  However, for some vessel owners/operators the benefit 
of in-service maintenance is pe rceived as marginal relative to  the costs, particularly for 
recreational vessels (Hilliard et  al. 2006).  In addition, where macroalgal growth has been 
allowed other species, such as mobile invertebrates, are likely to utilise the habitat the macroalgae has created.  As a result, macr oalgae may facilitate the entry of other non-
indigenous species.   Some macroalgal species may survive on vessel hulls  in life stages that are difficult to remove 
and/or are not visible, for example U. pinnatifida  (Coutts 1999).  In addition, macroalgae can 
become entangled in hull appendages which, if  occurring on departure, may not be removed 
until the vessel reaches its destination.  C. taxifolia , for example, has i nvaded estuarine areas 
of New South Wales where recreationa l vessels are known to moor.  As C. taxifolia  can 
regenerate from fragments, the clearing of entangled fragments within the arrival port may 
110  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
lead to establishment.  Other species, such as C. fragile  ssp. fragile,  may also be translocated 
in this manner. 
 Many algal species have been known to have wide  tolerances to envir onmental stressors.  For 
example, the invasive Gracilaria salicornia  can withstand both high and low temperatures 
and salinities, and U. pinnatifida  is known to be able to withstand exposure environments 
ranging from sheltered harbours to  open coasts (Russell et al. 2008) .  To survive translocation 
macroalgal species need to be able to tole rate a range of conditions which may include 
biocides from anti-fouling paint, shear forces , temperature variation and nutrient variation.  
Many macroalgal species have been found to tole rate most, or all, of these factors.   
 Given that particular macroalgal species: 
 are known to occur along vessel pathways; 
 are known to be associated with vessel biofouling; 
 are known to survive translocation; 
it is concluded that the likelihood of entry of  non-indigenous macroalg al species into New 
Zealand via vessel biofou ling is non-negligible. 
 
4.12.2.2.  Establishment assessment 
The m
acroalgal species arriving as hull fouling ar e likely to be tolerant to a wide variety of 
environmental stressors (e.g. Ulva  spp. Heesch et al. 2007, 2009).  As a result, it is assumed 
that most species surviving tr anslocation to New Zealand have  the potential to settle and 
establish, either as fragments or gametes.  Ho wever, the long-term viability of the population 
will depend upon the species ability to adapt to the local conditions and compete against the 
local flora. 
 The ability of a non-indigenous species to compete for sheltered space in a disturbed 
environment is key to establishment success (M ineur et al. 2008).  Moreover, it has been 
noted that invasive species are competitively su perior on artificial substr ates such as marina 
pontoons (Bulleri et al. 2006, Farrell & Fletcher 2006).  As a consequence, ports may provide 
‘stepping stones’ for macroalgal species to esta blish in New Zealand (Floerl et al. 2009).  For 
example, the major ports of first arrival for recreational vessels (Opua and Auckland) have 
substantial areas occupied by marina and moor ing structures as well as many vessels which 
may be susceptible to fouling.  From such stepping stones non-indigenous species can take advantage of disturbance generated gaps in the natural assemblages to colonise natural 
substrates.  C. fragile  ssp. fragile  has been reported to colonise  such gaps with dense stands 
which prevent the natural assemblage re-establishing (Schei bling & Gagnon 2006). 
 The attributes of invasive macroalgae have been comprehensively reviewed (Williams & Smith 2007).  While the three macroalgal clas ses all have representatives known to be 
invasive, some genera have characteristics which predispose them to being invasive.  For example, successful green macroalgae are from the genera Codium , Caulerpa and Bryopsis  
which are characterised by their siphonous cons truction and tolerance to a broad range of 
habitats.  Of the families Williams & Smith (2007) identified as having significantly more invasive species than expected only the Ce ramiaceae are substantially represented in the 
results of the MAFBNZ commissioned research  on biofouling (Inglis et al. in prep.).  
However, the majority of species found conformed to the functional gr ouping of ‘filamentous’ 
which contains species expected to be most successful where disturba nce potential is high, 
such as port and marina environments (Williams & Smith 2007).    
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  111  
112  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis In New Zealand ten species of Chlorophyta,  11 species of Ochrophyta and 14 species of 
Rhodophyta have been identified as e ither non-indigenous or cryptogenic14 and established in 
New Zealand (Kospartov et al. in press). 
 Given that particular macroalgal species: 
 are known to inhabit similar climates/environmen ts to those that exist in New Zealand; 
 are capable of forming establishi ng populations in New Zealand; 
 have been recently recorded as  established in New Zealand; 
it is concluded that the likelihood of establis hment of non-indigenous macroalgal species in 
New Zealand via vessel biof ouling is non-negligible. 
 
4.12.2.3.  Consequence assessment 
The im
pacts of introduced macroalgae are generally poorly understood, with Williams & 
Smith (2007) reporting impact studies for just 17 of the 277 macroalgal  species known to be 
introduced. From these impact studies, cha nges in community composition, structure and 
function have been reported (Battershill et al. 1998, Gribben & Wr ight 2006, Strong et al. 
2006, Antoniadou & Chintinaglou 2007, Wright  et al. 2007, Piazzi & Balata 2009).   
 
 The case of the invasive macroalgae species C. taxifolia has been well documented. In 1984, a 
cold tolerant and fast-growing strain of C. taxifolia  was accidentally released from the 
Oceanographic Museum of Monaco into the Me diterranean Sea. By 2000, it had spread to 
Mallorca, Elba, Sicily and Tunisia (Meinesz et al. 2001 In Galil et al. 2008). Although not 
initially introduced via hull fou ling, the hulls of fishing boats  and recreational craft were 
important vectors for the secondary spread of  this species (Knoep ffler-Peguy et al. 1985, 
Meinesz, 1992 In  Galil et al. 2008). 
 Areas supporting stands of native Mediterranean seagrasses have been smothered by 
C. taxifolia  (Zibrowius 1991 In NIMPIS 2010) which has had negative effects on shoot 
density and size of native seag rass (Ceccherelli & Cinelli, 1997 In NIMPIS 2010). In 
addition, the toxic secondary metabolites produced by C. taxifolia  prevents it being grazed by 
macro-herbivores and inhibits the gr owth of native m acroalgae (Galil 2000 In NIMPIS 2010). 
As a result of this protective mechanism, the remaining native species may suffer from 
overgrazing (Boudouresque 1997  In NIMPIS 2010). In addition, C. taxifolia  limits the light 
levels to benthic species, and traps se diments and organic matter (Boudouresque 1997 In 
NIMPIS 2010).  As a result of the aforemen tioned effects, the invasive C. taxifolia  has lead to the 
disappearance of encrusting algae and reduced di versity of associated benthic assemblages 
with impacts observed on populations of fi sh, amphipods, sea urchins and polychaetes 
(Boudouresque et al. 1995 In NIMPIS 2010). 
 Reduction of available fish habitat has reduced catches. In addition, fish that can consume C. 
taxifolia , such as the Mediterranean bream (Sarpa salpa), are rendered unsuitable for human 
consumption by their accumulation of toxins (Meinesz & Hesse 1991 In GISD 2006). 
Decreases in fishing efficiency have been cause d by entanglements in boat anchors, fishing 
nets and trawling gear (NIMPIS 2010).   
                                                
 
14 Cryptogenic species are defined as those species for which the native range is ambiguous.  This is usually due 
to a lack of information about the speci es or they are new to science or they have been present in New Zealand 
for a long time (Cranfield et al. 1998, Kospartov et al. 2008).  
In addition, the introduction of  macroalgae can alter social values, for example, the 
introduction of C. taxifolia  has decreased the aesthetic value of underwater habitats and 
reducing tourism and scuba diving activities (M einesz 1999). In the State of Hawai’i, the 
island of Maui contends with a US$20 million yearly loss of tourism revenue due to the 
fouling of beaches by non-indigeno us algae (Beck et al. 2008).  
 In New Zealand, the seagrass habitats of the Kaipara harbour are essential for the snapper 
fishery of the west coast of the North Island as they provides habitat for approximately 90% of juvenile fish (Morrison 2008).  Given that sea grass habitats  have been overgrown by the 
invasive strain of C. taxifolia  (Williams & Smith 2007), such Ne w Zealand habitats are likely 
to be at risk along with th eir associated ecosystem serv ices including fisheries.   
 Economic impacts resulting from the cost of Caulerpa spp. eradication included 
approximately $US6 million spent in S outhern California up to 2004 (Anderson, 2004 In 
GISD 2006) and $AUS6-8 million in South Australia (GISD 2006).  Many native macroalgae provide the structure,  habitat and food which drive many coastal 
ecosystems.  As a result, any modification to  the species composition and abundance is likely 
to have flow-on effects to coastal marine values.    
4.12.3.  Risk estimation 
Given that particular m
acroalgal species: 
 are known to be associated with vessel biofouling; 
 may facilitate the entry of a variety of other non-indigenous organisms; 
 are capable of forming establishi ng populations in New Zealand; 
 have well known invasive behaviours; 
 where introduced, have been a ssociated with high impacts; 
 have the potential impact to disturb the na tive macroalgal structure which may have far 
reaching indirect impacts to core values; 
it is concluded that the impor t of non-indigenous macroalgal species via biofouling poses a 
non-negligible risk to New Zealand’s marine core values. 4.12.4.  Risk management 
A generic ap 
proach to the rem oval of biofouling would be the appropriate risk management 
option of macroalgal hull fouling (Section 3).  Key considerations for macroalgal risk 
management during mitigation actions are the containment (and appropriate neutralisation) of 
macroalgal fragments, including reproductive structures (such as spores, gametophytes), and any associated mobile organisms.  
4.12.5.  References 
Ada 
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Marine Biological Association of the United Kingdom  87: 629-641. 
Battershill, C; Miller, K; Cole, R (1998) The understorey of marine invasions.  Seafood New 
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Beck, KG; Zimmerman, K; Schardt, JD; Stone, J; Lukens, R R; Reichard, S; Randall J; 
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Callow, M E (1986a) Fouling algae from in-service ships.  Botanica Marina  29:351-357. 
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Coutts, A D M (1999) Hull fouling as a modern vector for marine biological invasions: 
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Scheibling, R E; Gagnon, P (2006) Co mpetitive interactions between the invasive green alga 
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4.13. MICROALGAE 
4.13.1.  Hazard identification 
Microalgae are a taxonom
ically di verse group of unicellular algae which can be found in the 
plankton or on the benthos, si ngularly or in aggregations, as free-living, commensal or 
symbiotic organisms.  As primary producers, mi croalgae are important components of marine 
ecosystems providing up to 40% of global primary production (Falkowski 1994). In addition, 
microalgae link dissolved nutrients to macro-or ganisms in the zooplankton and on, or in, the 
benthos.  Some microalgae can form blooms with  harmful or negative consequences, such as 
impacts on human health and other values rela ted to the marine environment.  In New 
Zealand Alexandrium  spp. and Gymnodinium  spp. have caused outbreaks of shellfish 
poisoning which have had significant impacts on aquaculture operations, recreational and cultural harvests and human health (MacKenzie et al. 1997, Mackenzie & Beauchamp 2001).  
However, while toxic bloom events  have been relatively recent  occurrences in New Zealand it 
is difficult to determine if the causative  species are native or non-indigenous. 
 Globally it has been difficult to determine if microalgal species are introduced due to 
taxonomic difficulties and/or a lack of ba seline information (Bolch & Reynolds 2002, 
Lundholm & Moestrup 2006).  Those microalgae th at have been identified as non-indigenous 
species in the literature tend to be species dete cted due to negative bloom events, for example, 
Gymnodinium catenatum, Thecadinium  yashimaense , and Karenia  mikimotoi  (Hallegraeff & 
Bolch 1992, McMinn et al. 1997, H oppenrath et al. 2007).  Ho wever, many microalgal 
species are known to be carried in ballast water (Hallegraeff & Bolch 1992) and a link 
between shipping and invasion of Tasmanian waters by G. catenatum  has been shown 
(McMinn et al. 1997).  Whilst all documented non- indigenous microalgae are thought to have 
been translocated by ballast water or with aquaculture spec ies (Hallegraeff & Gollasch 2006), 
microalgae are also well know n components of biofilms. 
 Microalgae associated with marine biofilms are typically diatoms of Navicula spp. 
Achnanthes spp. Amphora spp. and Nitzschia  spp. (Callow 1986, Anil et al. 2006) which are 
generally considered innocuous in  terms of potential impacts.  While species of these genera 
normally dominate biofilms they are not necessarily the only microalgae fouling the surface.  Maso et al. (2003), for example, have identifi ed harmful algal bloom species (dinoflagellates 
of Ostreopsis , Coolia  and Alexandrium  spp.) adhering to the surfa ce of plastic debris in the 
Mediterranean. This finding flags the possibili ty that hull biofouling has been an overlooked 
vector for the translocation of microalgal sp ecies.  Similarly, Maso et al. (2003) found both 
vegetative and encysted forms of the dinoflage llates of concern adhering to debris. In 
addition, G. catenatum  cysts were found to be fouling oyster sticks during the 2000/2001 
bloom in the northwest of the North Island of  New Zealand (Mackenzie & Beauchamp 2001).   
 The cyst stages of m
any harmful algal specie s are known to be resistant to a variety of 
stressors, thus cysts adhered to vessel hulls may not be vulnerable to treatments other than 
physical removal.  Cysts do not re quire light and therefore are not  restricted, as are vegetative 
cells, to areas of the hull near the waterline wher e shear stress and abrasion are likely to lead 
to dislodgement. Therefore, cysts could be lodged in niche areas and consequently be 
protected.  Microalgae have been identified as compone nts of hull biofilms (Callow 1986, Cooksey & 
Wigglesworth-Cooksey 1995, Callow & Callow 2002, Cassé & Swain 2006).  However, the 
identified species are usually restricted to  dominant, commonly identified and reportedly 
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  117  
innocuous species.  Biofilm species were not  targeted in the MAFBNZ commissioned 
research on biofouling, however the few diatom samples collected were only able to be 
identified as Bacillariaceae and Melosirai (Inglis et al. in prep.) .  However, given the 
ubiquitous presence of biofilms on submerged surf aces, it is likely that microalgae are present 
on the hulls of most vessels arriving in New Zealand.   
 Given that microalgae: 
 are known components of marine  biofouling assemblages;  
 are known to have been introduced to new locations; 
 have known impacts on core values; 
it is concluded that microalgal species repr esent a potential hazard via vessel biofouling. 
 
As current anti-fouling system t echnologies only have the capacity  to control biofouling to the 
biofilm (slime layer) level, the prevention of  the entry of non-indigenous microalgae via 
vessel hull fouling cannot be practically achiev ed. As a consequence, further research is 
required to confirm the presence and transl ocation potential of potentially hazardous 
microalgae before the development and implementation of biofilm control measures can be justified specifically from the biosecurity perspective.  
4.13.2.  References 
Anil, A C; Patil, J S; Mitbaukar, S; D
’Costa, P M; D’Silva, S; Hegde, S; Naik, R (2006) Role 
of diatoms in marine biofouling. In Tewari, A (Ed) Recent advances on applied 
aspects of Indian marine algae w ith reference to global scenario . Volume 1. Central 
Salt and Marine Chemicals Institute. pp. 351-365. 
Bolch, C J S; Reynolds, M J (2002) Species  resolution and global distribution of 
microreticulate dinoflagellate cysts. Journal of Plankton Research  24: 565-578. 
Callow, M E (1986) A worldwide survey of slim e formation on anti-fouling paints. In Evans, 
L V; Hoagland, K D (Eds) Algal Biofouling.  Studies in Environmental Science  28. 
Elsevier; Amsterdam, The Netherlands. 
Callow, M E; Callow, J A (2002) Mari ne biofouling: a sticky problem. Biologist  49: 1-5. 
Cassé, F; Swain, G W (2006) The development of micro-fouling on f our commercial anti-
fouling coatings under static and dynamic immersion.  International Biodeterioration 
and Biodegradation  57: 179-185. 
Cooksey, K E; Wigglesworth-Cooksey, B (1995)  Adhesion of bacteria and diatoms to 
surfaces in the sea: a review. Aquatic Microbial Ecology  9: 87-96. 
Falkowski, P G (1994) The role  of phytoplankton photosynthesis  in global biogeochemical 
cycles. Photosynthesis Research  39: 235-258. 
Hallegraeff, G M; Bolch, C J (1992) Transport of diatom and dinoflagellate resting spores in 
ships’ ballast water: implications fo r plankton biogeography and aquaculture. Journal 
of Plankton Research  14: 1067-1084. 
Hallegraeff, G; Gollasch, S (2006) Anthropogenic introductions of microalgae. In Granéli, E; 
Turner, T (Eds) Ecology of Harmful Algae . Ecological Studies 189. Springer Verlag; 
Berlin, Germany;  pp. 379-390. 
Hoppenrath, M; Elbrächter, M; Halliger, H;  Koeman, R P T; Kra khmalnyy, A; Surek, B; 
Erler, K; Luckas, B (2007) First record s of the benthic, bloom-forming, non-toxic 
dinoflagellate Thecadinium yashimaense (Dinophyceae) in Europe: with special 
emphasis on the invasion in the North Sea. Helgoland Marine Research  61: 157–165. 
Inglis, G J; Floerl, O; Unwi n, M; Ponder-Sutton, A; Seawar d, K; Kospartov, M; Bell, A; 
Kluza, D (in prep.). The biosecurity risks associated with biofouling on international 
118  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  119 vessels arriving in New Zealand. Summary of the patterns and predictors of fouling.  
Biosecurity New Zealand Technical Paper. 
Lundholm, N; Moestrup, Ø (2006) The biogeography of  harmful algae. In Granéli, E; Turner, 
J T (Eds) Ecology of Harmful Algae . Ecological Studies Vol 189. Springer-Verlag; 
Berlin, Germany; pp 23-35. 
Mackenzie L; Beauchamp T (2001) Gymnodinium catenatum  in New Zealand: a new problem 
for public health and the shellfish in dustry. Website publication, The Cawthron 
Institute www.cawthron.org.nz. 
MacKenzie, L; Adamson, J; Todd, K (1997) Alexandrium  spp. and associated PSP-toxin 
contamination of shellfish on the No rth Island east coast during 1997. In Proceedings 
of the marine biotoxin science workshop No. 8 , MAF Regulatory Authority; 
Wellington, New Zealand; pp 58-65. 
McMinn, A; Hallegraeff, G F; Thomson, P; Jenkinson, A V; Heijnis, H (1997) Cyst and 
radionucleotide evidence for the recent introduction of the toxic dinoflagellate 
Gymnodinium catenatum  into Tasmanian waters. Marine Ecology Progress Series  
161: 165-172. 
Maso, M; Garcés, E; Pagés, F; Camp, J (2003) Dr ifting plastic debris as a potential vector for 
dispersing Harmful Algal Bloom (HAB) species.  Scientia Marina  67: 107-111. 
 
       
 
    
4.14. OTHER ARTHROPODS 
4.14.1.  Hazard Identification 
Phylum
 Arthropoda consists of two major extant  taxa - Chelicerata (e.g. horseshoe crabs, sea 
spiders, and arachnids), and Mandibulata. Mandi bulata contains the classes Crustacea (e.g. 
crabs, barnacles, water fleas) and the Trachea ta (e.g. insects, centipedes, millipedes). These 
taxa share several characteristi cs, including; a segmented body, a chitinous exoskeleton with 
ecdysis (moulting), jointed and paired segmental appendages, and the absence of locomotary 
cilia (Ruppert et al. 2004). For the purposes of  this hazard identification the following groups 
have been excluded from consideration under th e term “other arthropods”: the Tracheata as 
they are primarily terrestrial species and w ithin Crustacea, amphipods and isopods, barnacles 
and true crabs have been examined separately. The classes within the scope of “other 
arthropods” include Merostomata, Ma lacostraca, Ostracoda and Pycnogonida. 
 The ‘other arthropods’ have a variety of lif e-histories. Their f eeding and reproductive 
methods are highly varied and most are relatively small to medium in size. 
 The majority of invasive species included within  ‘other arthropods’ have freshwater affinities. 
For example, non-indigenous species of crayfish (Niwa & Ohtaka 2006), small crustaceans such as the tanaid C ercopagis pengoi,  and the cladoceran Daphnia lumholtzi , are from 
planktonic freshwater/brackish habitats. Due to their habitat rest rictions it can be assumed that 
it is unlikely that these organisms will survive translocation through marine environments via 
vessel biofouling. Although resti ng egg stages are known, the risk  of their transport can be 
discounted as they develop in soft sediment a nd are sensitive to small temperature variations 
(Sopanen 2008).     There are few records of ‘other arthropods’ in hull fouling surveys. Gr oups that have been 
recorded during hull fouling su rveys include the Mysidacea, Tanaidacea, Calanoidea and the 
Chelicerate groups Pycnogonida and Xiphosura (Cranfield et  al. 1998, Cohen et al. 2001, 
Gollasch 2002, Cardigos et al. 2006, Lewis et  al. 2006, Coutts & Dodgshun 2007, Farrapeira 
et al. 2007). The MAFBNZ commissioned rese arch on biofouling found an indeterminate 
Tanaidacea on one vessel and the no n-indigenous unestablished pycnogonid Endeis flaccida  
on 7 vessels (Inglis et al. in prep.).   Given that “other arthropods”: 
 are rarely known as components of  vessel biofouling assemblages; 
 have few known introductions to new marine locations; 
 have no known impacts on core values; 
it is concluded that “other arthropod species” do not represent a potential hazard via vessel 
biofouling. 
  
4.14.2.  References 
Cardigos, F; Tem
pera, T; Avila, S; Goncal ves, J; Colaco, A; Santos, R S (2006) Non-
indigenous marine species of the Azores. Helgoland Marine Research  60: 160-169. 
Cohen, A N; Berry, H; Mills, C; Milne, D; Britton-Simmons, K; W onham, M; Secord, D; 
Barkas, J A; Bingham, B; Bookheim, B; By ers, J; Chapman, J W; Cordell, J; 
Dumbauld, B; Fukuyama, A; Harris, L H; Kohn, A; Li, K; Mumf ord, T; Radashevsky, 
120  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  121 V; Sewell, A; Welch, K (2001) Report of  the Washington State Exotics Expedition 
2000. Nearshore Habitat Program, Washington State Department  of Natural Resources. 
Coutts, A D M; Dodgshun, T J (2007) The nature a nd extent of organism s in sea-vessels: a 
protected mechanism for marine bioinvasions. Marine Pollution Bulletin  54: 875-886. 
Cranfield, H J; Gordon, D P; Willa n, R C; Marshall, B A; Batter shill, C N; Francis, M P; 
Nelson, W A; Glasby, C J; Read, G B (1998)  Adventive marine species in New 
Zealand. NIWA Technical Report 34. 48 pp. 
Farrapeira, C M R; Melo, A V de O M de; Barb osa, D F; Silva, K M E da (2007) Ship Hull 
Fouling in the Port of Recife, Pernambuco. Brazilian Journal of Oceanography  55(3): 
207- 221. 
Gollasch, S (2002) The importance of ship hull f ouling as a vector of species introductions 
into the North Sea. Biolfouling  18(2): 105-121. 
Inglis, G J; Floerl, O; Unwi n, M; Ponder-Sutton, A; Seawar d, K; Kospartov, M; Bell, A; 
Kluza, D (in prep.). The biosecurity risks associated with biofouling on international 
vessels arriving in New Zealand. Summary of the patterns and predictors of fouling.  
Biosecurity New Zealand Technical Paper. 
Lewis, P N; Bergstrom, D M; Whinam, J (2006)  Barging in: a temperate marine community 
travels to the subantarctic. Biological Invasions  8: 787-795. 
Niwa N, Ohtaka A (2006) Accidental introduction of symb ionts with freshwater shrimps . In 
Koike, F; Clout, M N; Kawamichi, M; De Poorter, M; Iwatsuki, K (Eds) Assessment 
and Control of Biological Invasion Risks.  Shoukadoh Book Sellers; Kyoto, Japan; and, 
IUCN; Gland, Switzerland; pp 182-186. 
Ruppert, E E; Fox, R S; Barnes, R D (2004) Inve rtebrate Zoology: a functional evolutionary 
approach (7th Ed). Brooks/Cole-Thomson Learning, Belmont, USA. 
Sopanen, S (2008) The effect of temperature on the development and hatching of resting eggs 
of non-indigenous predatory cladoceran Cercopagis pengoi  in the Gulf of Finland, 
Baltic Sea. Marine Biology 154: 99-108.  
4.15. PEANUT WORMS (PHYLUM: SIPUNCULA) 
4.15.1.  Hazard Identification 
Peanut worm
s (sipunculans) are unsegmented benthic worm-like organisms, related to 
annelids, molluscs and echiurans (Ruppert et  al. 2004). There are approximately 370 known 
sipunculan species that range from  the shallow subtidal habitats to the abyssal plains (Murina 
1984, Flach and Heip 1996, Ramire z-Llodra et al. 2008). Although the majority of species 
inhabit protective crevices (the substratum, d ead barnacles or echinoderm tests), others bore 
tubes into limestone, or are sediment dwe llers (Ruppert et al. 2004). Peanut worms are 
normally rare, however, they can in some ci rcumstances be highly abundant, for example, 
averaging 8,000 individuals per m2 on coral reefs (Chapman 1955 In Murina 1984).  
 
Ecologically, peanut worms feed non-selectively on particulate organic matter from the water 
column or on the benthos (Ruppert et al. 2004).  They have muscular bodies that lack any 
major calcareous formations, shell, or tube (Murina 1984). In addition to being a food source 
for Asian populations, due to their unprotected nature peanut worms are prey items for molluscs, actinians, crabs, sea stars, and fish (Kohn 1975 In Murina 1984, Nelson 1981 In 
Murina 1984). Peanut worms are proposed to be important bioeroders on coral reefs (Rice and 
McIntyre 1972 In Murina 1984, Rice 1976 In  Murina 1984, Wood 1999). 
 The peanut worm life-cycle involves a pelagi c larval phase and a sedentary adult phase. 
Almost all species have separate sexes, al though sexual dimorphism is not obvious (Cutler 
2009). Gametes are broadcast spawned into the wa ter column where they fertilise, disperse 
and develop (Cutler 2009).  Penaut worms are not common members of hull fouling communities and there have been no 
recorded introductions of these organisms. Howe ver, two individual sip uncula (identified to 
phylum level) were recorded in seachest s by Coutts and Dodgshun (2007), while in the 
MAFBNZ commissioned research on biofouling,  Inglis et al. (in prep.) identified one 
sipunculan (identified to phylum level). This lone specimen was sampled from the keel of a 
recreational vessel and associated  with heavy fouling consisting of predominantly barnacles. 
 The uncertainties and assumptions associated with the introduction and establishment of 
peanut worms via vessel biofouling include the general poor understa nding of sipunculan 
taxonomy, functional role, life-cycl e, and associations in the coastal marine environment 
(Cutler 2009).   Given that peanut worms: 
 are rarely associated with ma rine biofouling assemblages; 
 are not known to have be introduced to new locations; 
 have had no known effects on core values; 
it is concluded that peanut worms are not considered a po tential hazard.  
References 
Coutts, A D M; Dodgshun, T J (2007) The nature a nd extent of organism s in sea-vessels: a 
protected mechanism
 for marine bioinvasions.  Marine Pollution Bulletin  54:875-886. 
Cutler, E B (2009) Phylum Sipuncula. In Gordon, D P (ed) New Zealand Inventory of 
Biodiversity. Volume 1 : Kingdom Animalia; Radiata, L ophotrochozoa, Deuterostomia. 
Canterbury University Press; Ch ristchurch, New Zealand; pp 302-307. 
122  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  123 Flach, E; Hiep, C (1996) Vertical  distribution of macrozoobenthos  within the sediment on the 
continental slope of the Goba n Spur area (NE Atlantic).  Marine Ecology Progress 
Series  141:55-66. 
Inglis, G J; Floerl, O; Unwi n, M; Ponder-Sutton, A; Seawar d, K; Kospartov, M; Bell, A; 
Kluza, D (in prep.). The biosecurity risks associated with biofouling on international 
vessels arriving in New Zealand. Summary of the patterns and predictors of fouling.  
Biosecurity New Zealand Technical Paper. 
Murina, G V V (1984) Ecology of Sipuncula.  Marine Ecology Progress Series  17: 1-7. 
Ramirez-Llodra, E; Ballestero s, M; Company, J B; Dantart,  L; Sarda, F (2008) Spatio-
temporal variations in biomass  and abundance in bathyal non-crus tacean megafauna in 
the Catalan Sea (North-west Atlantic). Marine Biology 153: 297-309. 
Ruppert, E E; Fox, R S; Barnes, R D (2004) Invertebrate Zoology: A functional evolutionary 
approach (7th edition). Thompson Lear ning; Victoria, Australia. 
Smith, L D; Wonham, M J; McCann, L D; Ruiz , G M; Hines, A H; Carlton, J T (1999) 
Invasion pressure to a ballast -flooded estuary and an assessm ent of inoculant survival.  
Biological Invasions  1:67-87. 
Wood, R A (1999) Reef Evolution . Oxford University Press. 
 
  
4.16. ROUND WORMS (PHYLUM: NEMATODA) 
4.16.1.  Hazard Identification 
Round worm
s (nematodes), are among the most widespread and abundant animals on earth. 
They inhabit moist interstitial e nvironments in all habitats, in cluding the bodies of plants and 
animals. In the marine benthos, free-livi ng round worms are the most abundant animals 
present (Warwick 1981), for example, Jensen  (1987) found 2.9 million individuals per m2 
down to a 20 cm depth.  
 Ecologically, despite being ubiquitous and a bundant benthic meiofauna the role of round 
worms and their importance to food chains  remains unestablished (Heip et al. 1985 In 
Vranken et al. 1988, Herman & Vranken 1988). Free-living round worms feed on a range of organisms and particulate organic matter, al though, many are specialis ed feeders. Marine 
round worms may be carnivores, herbivores or detritivores, feeding on, for example, other round worms, diatoms, algae, fungi, bacteria, an d particulate organic matter (Ruppert et al. 
2004).   Reproduction in marine round worms varies w ith temperature, salin ity and quality and 
quantity of food. Under optimal conditions short- lived species are able to produce tens of 
generations per year (Tietjen & Lee 1972, Warwick 1981, Vranken et al. 1988). Round 
worms usually have separate sexes and repr oduce by internal fertili sation (Ruppert et al. 
2004).   The only identified invasive/introduced round worm recorded is Anguillicola crassus  which 
was introduced to Germany and Ireland. However, as an eel parasite, th is round worm is not 
associated with organisms of the fouling comm unity or thought to be vectored by hull fouling 
(Gollasch & Nehring 2006, Minchin 2007). While parasitic and terrestrial round worms have 
been recorded as having impacts on organisms and plant commodities, no impacts have been recorded for free-living marine round worms.  Records of round worms on hull fouling are scarce. Records include Oncholaimus oxyuris 
found on container vessels in th e North Sea (Gollasch 2002), an unidentified species found in 
a seachest in New Zealand (Coutts & Dodgs hun 2007), and unidentified individuals on 
fishing, research, tug, pilot and ca rgo vessels in the Port of Reci fe (Brazil) (Farrapeira et al. 
2007). In the MAFBNZ commissioned rese arch on biofouling, no round worms were 
recorded on the hulls of vessels sampled ente ring New Zealand (Inglis et al. in prep.). 
  Given that round worms: 
 are rarely known as components of  vessel biofouling assemblages; 
 have few known introductions to new locations; 
 have no known impacts on core values; 
it is concluded that round worm species do not represent a potentia l hazard via vessel 
biofouling. 
 
4.16.2.  References 
Coutts, A D M; Dodgshun, T J (2007) The nature a nd extent of organism s in sea-vessels: a 
protected mechanism
 for marine bioinvasions. Marine Pollution Bulletin  54: 875-886. 
124  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  125 Farrapeira, C M R; Melo, A V de O M de; Barb osa, D F; Silva, K M E da (2007) Ship Hull 
Fouling in the Port of Recife, Pernambuco. Brazilian Journal of Oceanography  55(3): 
207- 221. 
Gollasch, S (2002) The Importance of Ship Hull F ouling as a Vector of Species Introductions 
into the North Sea. Biofouling 18(2): 105-121.  
Gollasch, S; Nehring, S (2006) National checkli st for aquatic alien species in Germany. 
Aquatic Invasions  1(4): 245-269. 
Inglis, G J; Floerl, O; Unwi n, M; Ponder-Sutton, A; Seawar d, K; Kospartov, M; Bell, A; 
Kluza, D (in prep.). The biosecurity risks associated with biofouling on international 
vessels arriving in New Zealand. Summary of the patterns and predictors of fouling.  
Biosecurity New Zealand Technical Paper. 
Herman, P M J; Vranken, G (1988) Studies of th e life-history and energetics of marine and 
brackish-water nematodes. II. Produc tion, respiration and food uptake by Monhystera 
disjuncta . Oecologia  77: 457-463. 
Jensen, P (1987) Differences in microhabita t, abundance, biomass and body size between 
oxybiotic and thiobiotic free- living marine nematodes. Oecologia 71: 564-567. 
Minchin, D (2007) A checklist of alien a nd cryptogenic aquati c species in Ireland. Aquatic 
Invasions  2(4): 341-366. 
Ruppert, E E; Fox, R S; Barnes, R D (2004) Invertebrate Zoology: A functional evolutionary 
approach (7th edition). Thompson Lear ning; Victoria, Australia. 
Tietjen, J H; Lee, J J (1972) Life-cycles of Ma rine Nematode – Influence of Temperature and 
Salinity on the Development of Monhystera denticulata  Timm. Oecologia  10: 167-176. 
Vranken, G; Herman, P M J; He ip, C (1988) Studies of the lif e-history and energetics of 
marine and brackish-water nematodes. I. Demography of Monhystera disjuncta  at 
different temperature and feeding conditions. Oecologia  77: 296-301. 
Warwick, R M (1981) The Influence of Temperature and Salinity on Energy Partitioning in 
the Marine Nematode Diplolaimelloides bruciei. Oecologia  51: 318-325. 
 
  
4.17. RIBBON WORMS (PHYLUM: NEMERTEA) 
4.17.1.  Hazard Identification 
Ribbon worm
s (nemerteans) are bilateral unseg mented worm-like organisms. The phylum 
comprises of approximately 1,500 species, the ma jority of which occupy marine habitats, 
however, freshwater and terr estrial species exist (Gibs on 1995, Thollesson and Norenburg 
2003, Ruppert et al. 2004). Marine ribbon worms are commonly found in sediments and in the 
crevices among shells, stones, holdfasts of al gae and sessile animals (Ruppert et al. 2004). 
Ribbon worms have few predators as they often contain toxins or noxious  chemicals (Carroll 
et al. 2003). 
 Ecologically, ribbon worms are ubiquitous, however , their role and importance is undefined. 
Ribbon worms are carnivores that feed on the soft parts of arthropods, annelids, molluscs, fish 
and are parasites of crab eggs (Wickham et  al. 1984; McDermott & Roe 1985, Torchin et al. 
2001, Ruppert et al. 2004). A unique feature of ribbon worms is their long, eversible 
proboscis that is used in feeding and burrowing (McDermott & Roe 1985, Turbeville & 
Ruppert 1985, Ruppert et al. 2004). In armed ribbon worms, the proboscis is everted to expose a stylet at the tip from  which a toxin is injected to  subdue prey (McDermott & Roe 
1985, Ruppert et al. 2004). The discharged proboscis of unarmed ribbon worms is used to immobilise prey by the release of sticky toxic se cretions as it is coiled around them (Ruppert 
et al. 2004).  Reproduction in ribbon worms can be either sexu al or asexual. Most ribbon worm species 
have separate sexes, aggregate before spawni ng and fertilise externa lly (Thiel & Junoy 2006). 
Fertilisation may involve free spawning or mu cous bundles, however, some species undergo 
internal fertilisation (Thiel & Junoy 2006). Many ribbon worms, especially large-bodied species, fragment when disturbed and subse quently regenerate (R uppert et al. 2004). 
 Ribbon worms are not common components of hull fouling communities and there have been 
no reported invasive/introduced ribbon worm or  recorded impacts on core values. Ribbon 
worms have been recorded on the hulls of various vessels in Brazil (F arrapeira et al. 2007) 
and in seachests of vessels in New Zealand (Coutts & Dodgshun 2007). Ribbon worm larvae were also found in the ballast and cargo tanks of  container ships in North America (Smith et 
al. 1999). However, no ribbon worms were record ed in the MAFBNZ commissioned research 
on biofouling (Inglis et al. in prep.).   Given that ribbon worms: 
 are rarely known as components of  vessel biofouling assemblages; 
 have few known introductions to new locations; 
 have no known impacts on core values; 
it is concluded that ribbon worm species do not represent a potentia l hazard via vessel 
biofouling. 
 
4.17.2.  References 
Carroll, S; McEvoy, E G; Gibson, R (2003) The pr oduction of tetrodotoxin-like substances by 
nem
ertean worms in conjunction with bacteria. Journal of Experiment al Marine Biology 
and Ecology  288: 51-63. 
126  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
Coutts, A D M; Dodgshun, T J (2007) The nature a nd extent of organism s in sea-vessels: a 
protected mechanism for marine bioinvasions. Marine Pollution Bulletin  54: 875-886. 
Farrapeira, C M R; Melo, A V de O M de; Barb osa, D F; Silva, K M E da (2007) Ship Hull 
Fouling in the Port of Recife, Pernambuco. Brazilian Journal of Oceanography  
55(3):207- 221. 
Gibson, R (1995) Nemertean genera and species of the world: an annotated checklist of 
original names and description citations, synonyms, current taxonomic status, habitats 
and recorded zoogeographic distribution. Journal of Natural History  29(2): 271-561. 
Inglis, G J; Floerl, O; Unwi n, M; Ponder-Sutton, A; Seawar d, K; Kospartov, M; Bell, A; 
Kluza, D (in prep.). The biosecurity risks associated with biofouling on international 
vessels arriving in New Zealand. Summary of the patterns and predictors of fouling.  
Biosecurity New Zealand Technical Paper. 
McDermott, J J; Roe, P (1985) Food, Feeding Behavior and Feeding Ecology of Nemerteans. 
American Zoologist  25: 113-125. 
Ruppert, E E; Fox, R S; Barnes, R D (2004) Invertebrate Zoology: A functional evolutionary 
approach (7th edition). Thompson Lear ning; Victoria, Australia. 
Smith, L D; Wonham, M J; McCann, L D; Ruiz , G M; Hines, A H; Carlton, J T (1999) 
Invasion pressure to a ballast-flooded estuar y and an assessment of inoculant survival. 
Biological Invasions  1: 67-87. 
Thiel, M; Junoy, J (2006) Mating behavior of  nemerteans: present knowledge and future 
directions. Journal of Natural History  40(15-16): 1021-1034. 
Thollesson, M; Norenburg, J L (2003) Ribbon wo rm relationships: a phylogeny of the phylum 
Nemertea. Proceedings of the Royal Society of London Series B  270: 407-415. 
Torchin, M E; Lafferty, K D; Kuri s, A M (2001) Release from pa rasites as natural enemies: 
increased performance of a globally introduced crab. Biological Invasions  3: 333-345. 
Turbeville, J M; Ruppert, E E (1985) Compara tive ultrastructure a nd the evolution of 
Nemertines. American Zoologist  25: 53-71. 
Wickham, D E; Roe, P; Kuris, A M (1984) Tr ansfer of nemertean e gg predators during host 
molting and copulation. Biological Bulletin  167: 331-338. 
 
       
4.18. SEA ANEMONES AND CORALS  (CLASS: ANTHOZOA) 
4.18.1.  Hazard Identification 
Anthozoa is the la
rgest class within phyl um Cnidaria, containing over 6,000 species of 
exclusively marine anemones and corals (Ruppert et  al. 2004). This class is split into several 
orders, including sea anemones (Actiniaria), ha rd corals (Scleractinia ), and black corals 
(Antipatharia) (Species 2000). Colonial co rals, solitary corals and sea anemones are 
ubiquitous in the marine environment ranging fr om shallow tropical seas to deep polar sea-
beds (Cairns 1995, Freiwald & Roberts 2005).   
 Unlike the other cnidarians, the anthozoan lif e-cycle does not have a medusa lifestage, 
therefore, the polyp alone unde rgoes sexual and asexual repr oduction (Ruppert al. 2004). 
Most corals and some sea anemones form ge netically identical cl ones through budding of new 
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  127  
polyps (asexual reproduction). Individual polyps may remain attached  to each other to form a 
colonial organism, or separate to form individual organisms (Karlson 1986, Ting & Geller 
2000). As a result, colonies are essentially a single genetic organism that can live for a 
hundred years or more (Dahan & Benayahu 1997). Species that broad cast spawn are often 
observed in stable, benign environments, while in more variable envi ronments species that 
brood are more common (Szmant 1986 In Vermeij et al. 2007).   
 
Ecologically, corals and sea anemones play impo rtant roles as benthic suspension feeders, 
space occupiers, and habitat formers. Tentacles on the polyp extend into the water column or along the substratum and capture food using st inging cells. Depending on  the species, food 
can range from plankton and particulate or ganic matter to crabs and fish (Anthony 1999, 
Smith & Gordon 2003, Lira et al. 2008). Some shallow water anthozoans harbour photosynthetic symbionts, typically zooxant hellae, which exude carbohydrates that are 
utilised by the host (Falkowski et al. 1984, Hoegh-Guldberg 1999). Most shallow water 
anthozoans participate in bot h phototrophic and heterotrophic feeding to varying degrees 
(Anthony & Fabricius 2000). Corals  and their calcium carbonate sk eletons provide habitat for 
many organisms and support commercial, subs istence and recreational fisheries in both 
shallow and deep waters (Bellwood et  al. 2004, Roberts et al. 2005, Stone 2006). 
 Known invasive anthozoa include Carijoa riisei , Tubastrea coccinea, Haliplanella lineata , 
and Oculina patagonica .  Typical invasive lif e history characteristic s exhibited include high 
tolerance towards abiotic factor s, high growth rate, early matu rity, localised recruitment of 
larvae, and high asexual repr oductive output (Gollasch & Reim ann-Zurneck 1996, Fine et al. 
2001, Glynn et al. 2008, Kahng et al. 2008, Lira et al. 2008, Rodolfo-M etalpa et al. 2008b). In 
addition, nuisance anthozoan species have, to some degree, similar aspects of their life 
history. For example, budding/m onoclonal stands can be seen in sea anemones such as 
Nematostella vectensis  in the northern hemisphere (Reitzel et al. 2008) and Actinia tenebrosa  
in the southern hemisphere (Ayre 1984).  Given that corals and sea anemones: 
 are known components of marine  biofouling assemblages; 
 are known to have been introduced to new locations; 
 have known impacts on core values; 
it is concluded that coral and sea anemone sp ecies represent a potential hazard via vessel 
biofouling.  
4.18.2.  Risk Assessment 
4.18.2.1.  Entry assessment 
Corals and sea anem
ones are components of benthic fouling assemblages globally. Some 
anthozoan species, such as H. lineata  and C. riisei , are found amongst port and harbour 
fouling in close proximity to commercial and recreational vessels (Elredge & Smith 2001, 
Grigg 2003).  
 Corals and sea anemones are known to be associat ed with vessel biofouling. In general, these 
organisms constitute a minor component of the hull fouling assemblages surveyed (Zibrowius 
1992, Lewis 2004), however, they have been found on all vessel types (Gollasch & Reimann-Zurneck 1996, Fenner 2001, Gollasch 2002, Godwin 2003, de Paula & Creed 2004, Coutts & Dodgshun 2007, Creed & de Paula 2007, Farrapeira et al. 2007). It has been suggested that anthozoa recruit to substrata th at have been submerged from be tween four months to a year 
(Gollasch & Reimann-Zurneck 1996, Creed & de  Paula 2007, Farrapeira et al. 2007). While 
128  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
this may be applicable to corals, the wandering nature of some sea anemones, their pedal disc 
and heavy mucous may enable attachment to  clean hulls. For example, in the MAFBNZ 
commissioned research on biofou ling, Inglis et al. (in prep.) found two unidentifiable and two 
local species of sea anemone on the relatively clean hulls of recreational vessels. 
 
Corals appear to be associated with vessels that have extended  periods of inactivity. The soft 
tissue of corals is susceptible to the abrasion caused by macro-fouling and sediment in areas 
of high water movement. As a result oil rigs and barges are the likel y vessels to present 
suitable substrates for coral colonisation. This a ppears to be the most likely scenario for the 
majority of Tubastrea coccinea introductions and in the ca se of the introduction of 
Balanophyllia elegans  to Hawai’i. Recreational and fishing vessels have been implicated in 
the introduction of Oculina patagonica to Algeria (Sartotetto et al . 2008), however , there is a 
high likelihood that these vessels have been subject to poor maintenance regimes and have 
spent extended periods idle. Th ese factors increase the like lihood of coral settlement and 
maturation on a vessel hull. However, temperate corals are unlikely to be associated with 
vessel hull fouling as they most frequently inhabit continental slopes, plateux, rises and 
seamounts. They are rare in ports and on coastlines due to factors related to these environments, such as abrasi on, and their life-histories. 
 Uncertainty exists with regard to how anti-foul ing systems affect corals and sea anemones. It 
is assumed that coral and sea anemone larvae are not likely to survive settlement on bare anti-
fouling paint. Therefore the mo st probable locations for succe ssful attachment are on niche 
areas that are already fouled. The pedal disc and mucous cr eated by adult sea anemones may 
afford some protection from anti-f ouling biocides. Therefore, if an individual were to settle on 
an exposed hull area, it may be ab le to migrate to a niche area.  
 
Coral and sea anemones are known to survive tran slocation. Some translocations have been 
recorded across relatively short distan ces, for example, the translocation of Sagartiogeton 
undatus  from the United Kingdom and Ireland to Ge rmany (Gollasch 2002). However, corals 
and sea anemones have been translocated over la rge distances, for example, the sea anemone 
Aiptasia pallida  was translocated from its native range  in the Carribean and eastern USA to 
the Port of Recife on the easte rn tip of Brazil (Farrapeira 2007). Other examples of long 
distance translocations include the coral B. elegans from California to Hawai’i (Hellberg 
1995, Godwin 2003), and the sea anemone H. luciae  being found on a vesse l hull in Germany 
(Gollasch and Reimann-Zurneck 1996).  
 The MAFBNZ commissioned research on biofouling found five sea anemones on vessels entering New Zealand (Inglis et al . in prep.). This research in cluded the first record of the 
unestablished tropical sea anemone Anemonia  cf manjano (Kospartov et al. in press). The four 
other native species identified were Anthopleura aureoradiata , Anthothoe albocincta , 
Diadume neozelanica , Oulactis mucosa .  
 Given that particular coral and sea anem
one species: 
 are known to occur along vessel pathways; 
 are known to be associated with vessel biofouling; 
 are known to survive translocation; 
it is concluded that the likelihood of entry of  non-indigenous coral and sea anemone species 
into New Zealand via vessel biofouling is non-negligible. 
 
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  129  
4.18.2.2.  Establisment assessment 
There are known suitable environm
ents and habita ts in New Zealand for the establishment of 
non-indigenous species of corals and sea anemones as these organisms are found throughout 
New Zealand from coastal waters to the aby ssal depths (Cairns 1995). Given that some non-
indigenous hull fouling corals and sea anemone s are associated with port and harbour 
environments, it is likely that they will find similar habitats upon en try into New Zealand.  
 Corals and sea anemones are able  to establish through fragmentation or the release of adults 
or larvae. Corals may become dislodged from hull fouling. However most corals, especially 
azooxanthellate species, grow at rates of millimetres per y ear, therefore, for a sizable 
fragment to be produced the organism would likely be seve ral years old. As a result, 
dislodgement of fragments large enough to surviv e and establish are unlikely to occur from a 
well maintained vessel.   The sea anemone, H. lineata,  has been translocated from its native Japan to numerous 
locations in both hemisphe res (Molina et al. 2009). H. lineata has many characteristics that 
have contributed to its success as an invasi ve species, including, a propensity to bud, a high 
tolerance to physical conditions a nd the ability to detach from substrata and float on the water 
surface (Ayre 1984, Gollasch & Reimann-Zurneck  1996, Ting & Geller 2000, Reitzel et al. 
2008). Detachment of sea anemones is likely a response to unfavourable conditions, such as, 
disturbances or changes to physical  conditions or spatial congestion. H. lineata  is the only 
non-indigenous sea anemone to known to have established in New Zealand.  
 There are no recent records of coral establishment in New Zealand. The two non-indigenous coral species, Hoplangia durotrix  and Tethocyathus cylindraceus,  established in New Zealand 
were introduced in the 1940’s, ar e restricted to the north/north eastern North Island and have 
had no recorded impacts (Cranfield et al. 1998). The native distributions of these corals are north-east Atlantic and western Atlantic, respecti vely (Kospartov et al. in press). Tropical and 
subtropical coral species are restricted to th e northern parts of New Zealand due to the 
presence of warmer currents (Cairns 1995, Laing & Chiswell 2003). Seasonal variations 
become more pronounced with increasing latitude, in addition, there is re duced irradiance and 
a greater supply of inorganic nutrients, particulate food (Muller-P arker & Davy 2001) and 
increased disturbance. These factors lead to d ecreased metabolic rates in tropical species and 
increased competition from macroalgae and othe r fouling organisms (Veron 1995, Fine et al. 
2001, Rodolfo-Metalpa et al. 2008a). Therefore, corals adapted to  temperate regions are slow-
growing azooxanthellate, and are often restri cted to sheltered depths and locations. 
 Given that particular coral and sea anemone species: 
 are known to inhabit similar climates/environmen ts to those that exist in New Zealand; 
 are capable of forming establishi ng populations in New Zealand; 
 have been recorded as es tablished in New Zealand; 
it is concluded that the likelihood of establis hment of non-indigenous coral and sea anemone 
species in New Zealand via vesse l biofouling is non-negligible. 
 
4.18.2.3.  Consequence assessment 
The invasive C. riise
i was first recorded in Hawai’i in 1972 (Evans et al. 1974 In Kahng & 
Grigg 2005), at this time it was considered to be an environmentally and economically benign 
species. However, C. riisei have smothered native black coral colonies Antipathes spp.. This 
has had negative impacts upon the US$30 million per annum black coral jewellery trade 
(Grigg 2003, Grigg 2004). With regard to the pr otection of New Zealand native black coral, 
especially those that grow in the Kermadec Islands and the shallow depths of Fjordland, 
130  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
temperature appears to be a significant protective factor as C. riisei  does not thrive below 
23°C (Kahng & Grigg 2005). 
 T. coccinea , native to the Indo-Pacific (Glynn et al . 2008), is commonly regarded as one of 
the most invasive corals and has spread thr ough the tropical western Atlantic (Fenner 2001, de 
Paula & Creed 2004, Creed & de Paula 2007). In New Zealand this coral is only found in the Kermadec Islands (corals are common but do not  form coral reefs) suggesting that, as 
elucidated in the establishment assessment, corals are not suited to temperate coastal 
environments. The Kermadec Islands are isol ated and highly exposed which influences 
organism and larval recruitment. The corals on these islands, are few in species, do not form 
reefs, and cover only 20 to 40% of substrate (F rancis & Nelson 2003).  It is likely that the 
sub-optimal conditions for T. coccinea  growth and competition from other species will buffer 
the impacts of any non-indigenous coral introductions.  As a result of the information reviewed, th e likelihood of tropical or  sub-tropical corals 
establishing and exhibiting signi ficant consequences to New Zealand’s core values is 
considered low.  Despite occasional high abundances  of invasive sea anemones, there is a deficiency of 
ecological studies regarding their impacts (Elr idge & Smith 2001, Molina et  al. 2009, Inglis et 
al. in press). The sea anemone, H. lineata, was introduced to New Zealand prior to 1983 and 
has since established populations in the Wa itemata Harbour, Motueka and Lyttleton Harbour 
(Cranfield et al. 1998, Inglis et al. in press). At present, ther e have been no reported impacts 
attributed to this invasion (I nglis et al. in press). For example, no significant impacts on the 
indigenous anemone A. aureoradiata has been recorded since the introduction of H. lineata, 
as these species have been observed to occur in  the same habitats and attached to similar 
substrata (Molina et al. 2009, Phillips pers. obs.) . Due to the prevalence of native sea 
anemone species and the lack of recorded imp acts of invasive sea anemones, it is deemed 
unlikely that non-indigenous anemones w ill have significant consequences on native 
populations.  Contact with stinging sea anemones can have si gnificant consequences to human health. For 
example, the toxins of the blue striped anemone  can cause severe reactions such as urticaria, 
haemorrhage, ulceration or a chronic rash (R ucks & Heasman 2006). A lik ely mode of contact 
between humans and sea anemones would be dur ing the clearance of fouling organisms from 
mussel buoys. However, the consequences of contact with sea anemones in this manner can be managed by the wearing of personal protectiv e equipment (Rucks & Heasman 2006). It is 
assumed that the introduction of non-indigenous  stinging sea anemones is not common given 
the lack of literature regarding impacts to hum an health. Therefore, it is unlikely that the 
establishment of common non-indigenous sea anem one species will lead to increased human 
health concerns.      
4.18.3.  Risk Estimation 
Given that particular coral species: 
 are not common com
ponents of vess el hull biofouling assemblages; 
 are only associated with fouling on inactive, unmaintained vessels; 
 are unlikely to establish in the majority of New Zealand’s marine environments;  
 are few in the number of introductio ns and invasive species recorded; 
 have not been recently recorded as introduced to New Zealand; 
 are not recorded as being responsible for impact s to core values relevant to New Zealand;  
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  131  
it is concluded that the import of non-indi genous coral species via biofouling poses a 
negligible risk to New Zeal and’s marine core values. 
 
Given that particular sea anemone species: 
 are not common components of vess el hull biofouling assemblages; 
 have few establishments recorded in New Zealand; 
 are not widely recorded as being res ponsible for impacts to core values;  
it is concluded that the impor t of non-indigenous sea anemone species via biofouling poses a 
negligible risk to New Zeal and’s marine core values. 
 
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Reitzel, A M; Darling, J A; Sullican, J C; Fi nnerty, J R (2008) Global population genetic 
structure of the starlet an em one Nematostella vectensi s: multiple introductions and 
implications for conservation policy. Biological Invasions . 10: 1197-1213. 
Roberts, J M; Brown, C J; Long, D; Bates,  C R (2005) Acoustic mapping using a multibeam 
echosounder reveals cold-water coral reefs and surrounding habitats. Coral Reefs  24: 
654-669.  
Rodolfo-Metalpa, R; Peirano, A;  Houlbreque, F; Abbate, M; Ferrier-Pages, C (2008a) Effects 
of temperature, light and heterotrophy on th e growth rate and budding of the temperate 
coral Cladocora caespitose . Coral Reefs  27: 17-25. 
Rodolfo-Metalpa, R; Reynaud, S; Allemand, D; Ferrier-Pages, C (2008b) Temporal and depth 
responses of two temperate corals, Cladocora caespitosa and Oculina patagonica , from 
the North Mediterranean Sea. Marine Ecology Progress Series . 369:103-114. 
Rucks, B; Heasman, K (2006) Blue Striped Sea Anemone: a safety issue for marine farms. 
New Zealand Aquaculture Magazine March/April 2006. 
http://www.nzaquaculture.co.nz/ . Accessed 16 January 2010. 
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Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  135 Ruppert, E E; Fox, R S; Barnes, R D (2004) Invertebrate Zoology: A functional evolutionary 
approach (7th edition). Thompson Lear ning; Victoria, Australia. 
Sartoretto, S; Harmelin, J-G; Bachet, F; Be jaoui, N; Lebrun, O; Zi browius, H (2008) The 
alien coral Oculina patagonica  De Angelis, 1908 (Cnidaria, Scleractinia) in Algeria and 
Tunisia. Aquatic Invasions  3(2): 173-180. 
Smith, F; Gordon, D (2003) Sessile Inverteb rates. In Andrew N; Francis M (2003) The Living 
Reef – the ecology of New Zealand’s rocky reefs . Craig Potton Publishing; Nelson, New 
Zealand; pp 80-91. 
Species 2000. Catalogue of Life: Dynamic Checklist . 
http://www.catalogueoflife.org/dynamic-ch ecklist/browse_taxa.php. Accessed 3 March 
2009. 
Stone, R P (2006) Coral habitat in  the Aleutian Islands of Al aska: depth distribution, fine 
scale species associations, and fisheries interactions. Coral Reefs  25: 229-238. 
Ting, J H; Geller, J B (2000) Clonal diversity in introduced populations of an Asian sea 
anemone in North America. Biological Invasions . 2: 23-32. 
Vermeij, M J A; Frade, P R; Jacinto, R I R; Debrot, A O; Bak, R P M (2007) Effects of 
reproductive mode  on habitat- related differences in popul ation structure of eight 
Caribbean coral species. Marine Ecology Progress Series  351: 91-102.  
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Veron, J E N (1995) Corals in space and time: the biogeography and evolution of the 
scleractinia . UNSW Press; S ydney, Australia. 
Ward, S (1995) Two patterns of energy allocation for growth, reproduction and lipid storage 
in the scleractinian coral Pocilopora damicornis. Coral Reefs  14: 87-90. 
Zibrowius, H (1992) Ongoing m odification of the Mediterranean fauna and flora by the 
establishment of exotic species. Mesogee.  51: 83-107.  
4.19. SEA SQUIRTS (CLASS: ASCIDIACEA) 
4.19.1.  Hazard Identification 
Sea squirts (Ascidians) are sessile, filter-f eeding organism
s that have a global marine 
distribution (Millar 1969).  Sea squirts are classified within  phylum Chordata as their tadpole-
like larvae possess a notochord (Ruppert et al. 2004). The generalised s ea squirt is a filter 
feeder. Water is drawn in th e inhalant siphon and over a “bra nchial basket” where food is 
trapped within a mucous net. The filtered water is then expelled via the exhalant siphon 
(Petersen 2007).  The majority of  sea squirts occur in shallow, sheltered waters attached to 
hard natural and artificial substr ata, however, soft sediment dw ellers and pelagic forms (salps) 
also exist (Lambert and Lambert 2003, R uppert et al. 2004, Davi s and Davis 2005).  
 The general sea squirt life-cycl e involves a short-lived planktonic larval phase and a sessile 
adult phase. Adult sea squirts may be solitary or colonial. While solitary forms each have their 
own test (protective covering), colonial form s may share the same test (Millar 1969, Ruppert 
et al. 2004). In addition, some colonies share a common exhalant  opening (Millar 1969, 
Ruppert et al. 2004). Solitary s ea squirts are usually spherical or cylindrical, while colonial 
sea squirts often form mats or hanging aggreg ations. Colonial sea squirts can often survive 
fragmentation and form new coloni es (Bullard et al. 2007a).   
 Sea squirts are hermaphroditic and in general, solitary species are broadcast spawners while 
colonial species fertilise eggs internally  (Lambert 2005, Kott and Page 2007). Sea squirt 
larvae are non-feeding, tadpole-like and free-s wimming. The larvae can swim for a few hours 
before settlement on an appropriate substrate where they metamorphose into the sessile adult form (Lambert 2005, Kott and Page 2007). Sea squi rt dispersal is very limited under natural 
conditions due to the short time the larvae spend in the water column.  
 
Ecologically, sea squirts are important suspension feeders, capable of filtering vast quantities of water for the removal of plankton and organi c material (Kott et al . 2009). Sea squirts are 
also major space occupiers on the benthos and in the plankton (Kott et al. 2009). Their ability to bud enables them to rapidly cover available space on the ground or in the water column (Kott et al. 2009). Sea squirts ha ve been recorded from the trop ics to the sub-Antarctic (Primo 
and Vazquez 2008).  There are many instances of s ea squirt introductions, particul arly in North America and 
Europe. Impacts include decrease d local biodiversity (Blum et al. 2007) and ne gative effects 
on fisheries and aquaculture (Thompson and MacNair 2004, Davis and Davis 2006, Howes et 
al. 2007, Locke et al. 2007, Valentine et al. 2007b). Many non-indigenous sea squirts experience competitor and predator releas e in introduced areas, allowing significant 
aggregations to develop (Locke et al. 2007, Osman and Whitlat ch 2007).  They often become 
gregarious during summer, as temperatures a nd nutrients increase, a nd can smother sessile 
benthic structures and organisms (Bullard et al. 2007b, Howes et al. 2007). Well known invasive species include Botrylloides violaceus , Diplosoma listerianum , Stylea clava , 
Ascidiella aspersa , Didemnum  sp, Botryllus schlosseri, Ciona intesinalis . In New Zealand the 
introduced S. clava, D. vexillium , A. aspersa , and Eudistoma elongatum  have become 
nuisance or pest species. Non- indigenous sea squirts also faci litate the introduc tion of their 
epibionts, such as amphipods (Howey 1998, Lambert & Lambert 1998, Wyatt et al. 2005). 
 Given that sea squirts: 
 are known components of marine  biofouling assemblages;  
136  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
 are known to have been introduced to new locations; 
 have known impacts on core values; 
it is concluded that sea squirt  species represent a potential  hazard via vessel biofouling. 
4.19.2.  Risk Assessment 
4.19.2.1.  Entry assessment 
Sea squirts are known to occur on en
try pathwa ys into New Zealand. The conditions found in 
port and marina environments provide highly su itable habitat for sea squirts including the 
provision of hard substrates in shallow water and sheltere d and disturbed environments 
(Locke et al. 2007, Tyrrell and Byers 2007). Sea s quirts have very limited natural dispersal 
ability, however, the increase of artificial structures such as aquaculture equipment and ocean 
going vessels has resulted in th e dispersal of sea squirts fa r beyond their natural boundaries 
(Lambert 2005, Minchin and Sides 2006). 
 
Sea squirts are known to have been transloc ated beyond their natura l boundaries via hull 
fouling (Lambert 2005, Wyatt et al. 2005, Fulle r 2007). Recreational vessels have been 
reported to have spread sea squirts, for example Didemnum  spp. to Ireland (Minchin and 
Sides 2006), S. clava  to Ireland and the German Bight (Minchin et al. 2006, Krone et al. 
2007), and an Atlantic clade of Clavelina lepadiformis  to the Mediterranean (Turon et al. 
2003). In the MAFBNZ commissioned research on biofouling, sea squirts were present on 5% 
of vessels sampled (Inglis et al. in prep.). Four non-indigenous species were recorded by 
Inglis et al. (in prep.), two of whic h are not established in New Zealand. 
 Sea squirts appear to be sensitive to anti-fouli ng systems. Hull fouling sea squirts often appear 
to be secondary foulers requiring the presence of other organism s to act either as a buffer, 
between them and the biocides, or as a s timulant for larval settlement (Edmondson 1944, 
Wood & Allen 1958). Typical of fouling species, sea squirts take advant age of failures anti-
fouling systems, and as a result are often found in niche areas (Edmondson 1944, Wood & Allen 1958). Lambert et al. (2005) states that the adult stages of a number of species are 
tolerant to heavy metals.   The ability of sea squirts to surv ive translocation is perhaps faci litated by their protective tests 
and mucous formation. In general, sea squi rts are not found on the hulls of fast moving 
vessels, however, they may survive in sheltered locations or on low speed voyages (Davis and Davis 2004, Coutts et al. in pres s). Sea squirts may have relativ ely high attachment strengths, 
for example, Coutts et al. (in press) found Asterocarpa humilis  remained attached at speeds of 
up to 18 knots. The ability of small colonial fragments to rapidly regrow means that, for 
certain species, the whole colony need not survive translocation.   Given that particular sea squirt species: 
 are known to occur along vessel pathways; 
 are known to be associated with vessel biofouling; 
 are known to survive translocation; 
it is concluded that the likelihood of entry of  non-indigenous sea squirt species into New 
Zealand via vessel biofou ling is non-negligible. 
 
4.19.2.2.  Establishment assessment 
There are known suitable environm
ents and habita ts in New Zealand for the establishment of 
sea squirts. Most hull fouling sea squirts are adapted to port /marina conditions and have a 
relatively high tolerance to environmenta l stressors (Lambert 2005). For example, S. plicata  
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  137  
can tolerate brackish waters ranging from 22–34 ppt salinity, and also  survive in polluted 
environments (Thiyagarajan & Qian 2003). S. clava has also been noted to tolerate a wide 
range of salinities (Kluza et al. 2006). Although cold temperatures may slow the growth or 
kill sea squirts, it should not be assumed that the warmer waters of northern New Zealand 
offer the best sites for establishment as species  are adapted to inhabit a variety of climates. 
For example, in northeastern North America  Didemnum  sp. are more successful in colder 
seasons and in deeper water than Botrylloides violaceus  (Dijkstra et al. 2007, McCarthy et al. 
2007). Non-indigenous sea squirts may also succeed  in conditions not thought to be suitable 
for establishment, for example, E. elongatum  from tropical/subtropical  eastern Australia has 
established in Northland, New Zealand, desp ite some conditions, such as flooding events, 
considered as being outside of its tolerance.  Hard substrata are essential for the majority of sea squirts. Artificial structures, such as port and marina structures provide good settlement s ubstrates that are shad ed (most larvae are 
negatively phototactic at settlement), offers litt le competition, and are relatively predator free 
(Lambert 2005). The availability of  artificial substrata was a crit ical factor in the successful 
invasion of S. clava  in the Gulf of St. Lawrence (Canada)  where natural hard substrates are 
scarce (Locke et al. 2007). Establishment on artificial substrates provides local sources of propagules to enable colonisati on of other areas, including th ose with natural substrates 
(Minchin and Sides 2006, Coutts an d Forrest 2007, Tyrell and Byers 2007). 
 Sea squirts can establish if colony fragments ar e dislodged or if fertilised larvae are released, 
settle, and reproduce. The gregarious nature of sea squirts may result in multiple individuals occupying the same vessel hull. However, sea squirt establishment does not necessarily 
depend on the presence of more than one individual. For example, solitary forms, such as Corella inflata , can self-fertilise (Lambe rt 1968) and colonial forms can asexually grow and 
disperse. In addition, non-id entical clones of colonial  sea squirts, such as Diplosoma 
listerianum , can effectively fuse /interbreed (Sommerfeldt and Bishop 1999). 
 Three non-indigenous sea squirt species have established in New Zealand since 2004 (Kospartov et al. in press).    Given that particular sea squirt species: 
 are known to inhabit similar climates/environmen ts to those that exist in New Zealand; 
 are capable of forming establishi ng populations in New Zealand; 
 have been recently recorded as  established in New Zealand; 
it is concluded that the likelihood of establis hment of non-indigenous sea squirt species in 
New Zealand via vessel biof ouling is non-negligible. 
 
4.19.2.3.  Consequence assessment 
In general, non-indigenous sea squirts can have ecological im
pacts by acting as larval 
predators and competing for resources leading to the reduction of growth  and survival of co-
occurring species. For example, the presence of S. clava was shown to reduce the settlement 
of other species, the exceptions being barn acles and spirobiid polychaetes (Osman and 
Whitlach 1999 In NIMPIS 2010). By using experimentally  manipulated settlement panels, 
Blum et al. (2007) examined the effects of non-indigenous C. intestinalis  on the sessile 
invertebrate community in San Francisco Bay, USA. They concluded that the presence of 
C. intestinalis  both depresses local species diversity a nd alters community assembly processes 
to change the community composition (Blum et al. 2007). 
 
138  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
D. vexillum has significant impacts on the species composition of benthic communities. For 
example, the abundance of two polychaete species, Nereis zonata and Harmothoe extenuata , 
increased significantly in infested areas compar ed with uninfested areas (Lengyel et al. 2009). 
By contrast, Mercer et al. (2009) found that D. vexillum mats had minimal influence on 
benthic macro-invertebrate species abundance and richness on pebble-cobble habitats in 
eastern Long Island Sound, New England, USA.  
 The establishment of non-i ndigenous sea squirts has re sulted in severe economic 
consequences to North American shellfish aquacu lture. In Nova Scotia, Canada, the sea squirt 
C. intestinalis  has caused up to 50% mort ality to blue mussels ( Mytilus edulis ) and incurs a 
loss of up to C$1.86 for every 1 kg of sea squirt  per m of mussel rope (Daigle & Herbinger 
2009). The clubbed tunicate ( S. clava ) has also been recorded to cause major economic 
impacts due to resource competition and smotheri ng resulting in reduced mussel production in 
Prince Edward Island, Canada (Davis & Davi s 2009). In addition to decreasing stock 
marketability and yields, sea squirt fouling ha s increased labour and processing costs. For 
example, for mussel aquaculture in Prince Edwa rd Island, Canada, these additional costs were 
conservatively estimated at 19% per pound of the market price (Thompson & McNair 2004). 
In north east North America, D. vexillum  has threatened the US$400 million sea scallop 
industry by having smothered up to 75% of surf ace area in the Georges Bank fishery (Morris 
et al. 2009). In addition, Atlantic cod, Atlantic herring, and haddock fisheries (US$55 million in 2004) have been negatively affected (Daley & Scavia 2008). Further,  dense fouling of sea 
squirts on fishing equipment, moorings and rope s can be time consuming to remove and can 
result in tangling of fishing gear.  Human health consequences within the mari culture industry have been related to the 
introduction of non-indigenous sea squirts. For example, asthma-like conditions have 
developed in workers cleaning S. clava fouled shellfish in poorly ventilated conditions (Kluza 
et al. 2006). In addition, soft tissu e injuries have resulted as wo rkers pulling in shellfish lines 
have to contend with the increased weight due to sea squirt fouling (Kluza et al. 2006). In 
some cases, workers are now exposed to caustic or harmful products used to manage sea 
squirt fouling. In addition, worker morale can also be negative ly affected due to the extra 
work involved in cleaning (Thompson & MacNair 2004).       In New Zealand D. vexillum  was originally recorded in Whangamata Harbour, northern North 
Island, but was introduced shortl y afterward to the Marlborough District, top of the South 
Island (Coutts & Forrest 2007). Give n the characteristics of the or ganism, the shift in habitat 
to cooler sheltered waters may have aided its establishment in New Zealand (Dijkstra et al. 
2007, McCarthy et al. 2007,  Valentine et al. 2007a). D. vexillum  fouls commercial mussel 
lines in the Marlborough Sounds, costing grow ers an estimated NZ$2.2 million per year in 
production losses (NZIER 2008). The Marlborough and Coromandel regions would likely be 
the most vulnerable aquaculture areas to sea sq uirt establishment as they produce 68 and 21% 
of New Zealand’s green lipped mussel expor ts, respectively (Aquaculture New Zealand 
2009). In 2008, the aquaculture industry was va lued at NZ$265 million, of which green lipped 
mussels made up 77% of exports by value.   The eradication, or m
itigation, of sea squirts is resource intensive. Solitary sea squirts can be 
prolific, for example, S. clava  can reach densities of 500 to 1,500 individuals per m2 (NIMPIS 
2010). Thus far, the New Zealand incursion response to S. clava has cost $2.2 million. 
Colonial sea squirts are particularly difficult to manage, for example, as D. vexillum  is highly 
fragmentary, 100% of the colony has to rem oved/contained or it can re-proliferate. In 
addition, D. vexillum  has highly variable growth forms allowing for the colonisation of a 
variety of substrata which can also create di fficulties for eradicati on (Coutts & Forrest 2007,  
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  139  
Pannell & Coutts 2007). In 2004, a preliminary eradication program at Shakespeare Bay, 
Marlborough, cost approximately NZ$0.65 million.  
 It is noted that in some cases S. clava  is being harvested and sold to Asian food markets at 
US$8 a pound. Although this is likely to be one of the only species able to be sold as it is 
large, relatively fleshy and solitary co mpared to colonial sea squirts. 
 
4.19.3.  Risk Estimation 
Given that particular sea squirts species: 
 are known to be associated with vessel biofouling; 
 m
ay facilitate the entry of a variety of other non-indigenous organisms; 
 are capable of forming establishi ng populations in New Zealand; 
 have well known invasive behaviours; 
 have had high impacts on core values following establishment; 
 are likely to have adverse effects on core values, such as, bivalve aquaculture and on 
coastal benthic communities; 
it is concluded that the impor t of non-indigenous sea squirts species via biofouling poses a 
non-negligible risk to New Zealand’s marine core values. 
4.19.4.  Risk Management 
The presence of sea squirts on hull fouling is  best treated with broad spectrum
 options 
(Section 3).  However, consideration should be given to the potential for the expulsion of 
propagules or fragmentation of colonies dur ing vessel haul out. Where identified, solitary 
forms may be removed by hand where it can be ce rtain all specimens are visible, it can be 
ensured that any stimulated spawning can be contained, and the removed organisms are 
contained for shore based disposal. Considera tions should also incl ude appropriate local 
containment due to the likelihood of release of  associated non-indigenous organisms into the 
receiving environment. 4.19.5.  References 
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of the invasive colonial tunicate Didemnum vexillum  on the recruitment of the bay 
scallop ( Argopecten irradians irradians ) and the  implications for recruitment of the 
recruitment of the sea scallop ( Placopecten magellanicus ) on Georges Bank. Journal of 
Aquatic Invasions  4(1): 201-211. 
NIMPIS (2010). Styela clava  impact details. National Introduced Marine Pest Information 
System (Eds: Hewitt C.L., Martin R.B., Sliwa C., McEnnulty, F.R., Murphy, N.E., Jones 
T. & Cooper, S.). Web publication <http://c rimp.marine.csiro.au/nimpis>, Accessed: 4th 
February 2010. 
NZIER (New Zealand Institute of Economic Research), 2008.  The Economic impact of 
Didemnum vexillum . Report to Marine Farming Association. 48 pp. 
Osman, R W; Whitlatch, R B (2007) Variation in the ability of Didemnum  sp. to invade 
established communities. Journal of Experimental Marine Biology and Ecology  342: 
40-53. 
142  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
Pannell, A; Coutts, A D M (2007) Tr eatment methods used to manage Didemnum vexillum  in 
New Zealand. A report prepared for Bios ecurity New Zealand. Biosecurity New 
Zealand, New Zealand; 43pp. 
Peterson, J K (2007) Ascidi an suspension feeding. Journal of Experimental Marine Biology 
and Ecology  342: 127-137. 
Primo, C; Vasquez, E (2008) Zoogeography of  the southern New Zealand, Tasman and 
southern African ascidian fauna. New Zealand Journal of Marine and Freshwater 
Research  42: 233-256. 
Ruppert, E E; Fox, R S; Barnes, R D (2004) Invertebrate Zoology: A functional evolutionary 
approach (7th edition). Thompson Lear ning; Victoria, Australia. 
Sommerfeldt, A D; Bishop J D D (1999) Random amplified polymorphic DNA (RAPD) 
analysis reveals extensive chimerism in a marine protochordate. Molecular Ecology  8: 
885-890. 
Sutherland, P (1978) F unctional roles of Schizoporella  and Styela  in the fouling community at 
Beaufort, North Carolina.  Ecology 59:257-264 
Thiyagarajan, V; Qian, P (2003) Effect of temperature, salinity and delayed attachment on 
development of the solitary ascidian Styela plicata  (Lesueur).  Journal of Experimental 
Marine Biology and Ecology 290:133–146 
Thompson, R; MacNair, N (2004) An overview of the clubbed tunicate ( Styela clava ) in 
Prince Edward Island. PEI Department of Agricultur e, Fisheries, Aquaculture and 
Forestry Technical Report 234 
Turon, X; Tarjuelo, I; Duran, S; Pascual, M (2003) Characterising i nvasion processes with 
genetic data: an Atlantic clade of Clavelina lepadiformis (Ascidiacea) introduction into 
Mediterranean harbours. Hydrobiologia  503:29-35 
Tyrell, M C; Byers, J E (2007) Do artificial substrates favour  non-indigenous fouling species 
over native species? Journal of Experimental Marine Biology and Ecology 342: 54-60. 
Valentine, P C; Carman, M R; Blackwood, D S; Heffron, E J (2007a) Ec ological observations 
on the colonial ascidian Didemnum s p. in a New England tide pool habitat. Journal of 
Experimental Marine Biology and Ecology  342: 109-121. 
Valentine, P C; Collie, J S; Ried, R N; Asc h, R G; Guida, V G; Blackwood, D S (2007b) The 
occurrence of the colonial ascidian Didemnum  sp. on Georges Bank gravel habitat – 
Ecological observations a nd potential effects on groundf ish and scallop fisheries . 
Journal of Experimental Marine Biology and Ecology 342: 179-181. 
Wood, E J F, Allen, F E (1958) Common marine fouling organisms of  Australian waters. 
Department of the Navy, Navy Office; Melbourne, Australia; 23 pp. 
Wyatt, A S J; Hewitt, C L; Walker, D I; War d, T J (2005) Marine intr oductions in the Shark 
Bay World Heritage Property, Western Australia: a preliminary assessment, Diversity 
and Distributions 11: 33-44. 
 
        
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  143  
4.20. SPONGES  (PHYLUM: PORIFERA) 
4.20.1.  Hazard Identification 
Sponges (phylum
 Porifera) are suspension f eeders commonly found in subtidal benthic 
habitats from tropical to polar  environments. Sponge habitat can range from the intertidal 
down to abyssal depths. Sponges can range in si ze and morphology from thin encrusting, or 
erect branching, to massive forms. Generally they become more dominant members of the 
benthic fauna as depth increases, hard substr ates become available and the water becomes 
relatively free from terreginous sediment and la rge salinity fluctuations.  Sponges are also 
abundant in high energy, subtidal environments where fast flowing currents can assist feeding 
efficiency, for example, seamounts (Rowden et al. 2002).   
 Sponges are important components of marine ecosystems. Sponges provide and facilitate the maintenance of biogenic habitats, couple th e benthic and pelagic environments, recycle 
nutrients, provide protection for other organism s within their matrix and are a source of 
primary production via their symbiotic and commensal relationships (Bell 2008).   
 Records of sponge introductions are few which is likely due to difficulties associated with 
their taxonomy.  For example, Kospartov et al. (in press) listed eight sponge species found in 
New Zealand waters as non-indigenous with a further 105 species lis ted as cryptogenic.  
Similarly, Streftaris et al. ( 2005) listed eight species of Po rifera as introduced to the 
Mediterranean, while Longo et al. (2007) described Paraleucilla magna as “the first record of 
an invasion due to Porifera for the Mediterranean Sea”.  There are few reported impacts from non-indi genous sponges.  Significant smothering 
impacts have been recorded as a result of the introduction of Mycale grandis to Hawai’i 
(Coles and Bolick 2007) and Terpios hoshinata  to Guam (Bryan, 1973, Rützler & Muzik 
1993). In addition, boring sponges (in particular Cliona spp.) have been shown to impact 
shellfish (Fromont et al. 2005).  Bioerosion, including shell bo ring, is a common attribute of 
sponge fauna and has been suggested as necessary to maintain coral reef  health (Hein & Risk 
1975).  Shell boring can have impacts on both th e productivity, and later marketability, of 
shellfish which can impact social, cultural and economic values.  For example, Fromont et al. 
(2005) reported that pearl oysters ( Pinctada maxima ) lost value, growth and resistance to 
predation when infected with boring species ( C. orientalis , C. dissimilis  and Pione velans).  In 
New Zealand, the non-indigenous boring sponge C. celata  has been collected from habitats 
ranging from the low intertidal to over 200 m in  depth with records da ting back to the mid-
1960’s (Probert et al. 1979).  While C. celata  is known to impact upon commercially 
important species such as paua and oysters it is generally considered a minor problem 
compared to boring Spionid bristleworms such as Polydora  spp.. 
 Sponges are known members of the fouling co mmunity and may be found on artificial 
substrates.  Sponges are not t ypically primary foulers and te nd to recruit later in the 
succession of organisms by using secondary meta bolites to clear and maintain space for 
growth.  While introductions have been attrib uted to hull fouling by some authors (Godwin 
2003) there have been few specimens recorded  by others. For example, in the MAFBNZ 
commissioned research on biofouling, the sol itary sponge sample recorded was the New 
Zealand native species Haliclonia  cf maxima  (Inglis et al. in prep.).   
 Given that sponges:  
 have a low likelihood of fouling vessels; 
 are likely to be associated with s ubstantial fouling by other species;  
144  Ministry of Agriculture and Forestry  Vessel Biofouling: Risk Analysis  
Ministry of Agriculture and Forestry Vessel Biofouling: Risk Analysis  145  are not typically invasive; 
 that are known to cause impacts, such as boring sponges, are already present in New 
Zealand; 
it is concluded that sponge species do not re present a potential haza rd via vessel biofouling. 
 
4.20.2.  References 
Bell, J J (2008) The functional role of m
arine sponges.  Estuarine, Coastal and Shelf Science  
79: 341-353. 
Bryan PG (1973) Growth rate , toxicity, and distributi on of the encrusting sponge Terpios  sp. 
(Hadromerida: Suberitidae) in Guam, Mariana Islands. Micronesica  9(2):237–242. 
Coles, S L; Bolick, H (2007) Invasive introduced sponge Mycale grandis  overgrows reef 
corals in K āne’ohe Bay, O’ahu, Hawai’i.  Coral Reefs  26:911 
Fromont, J; Craig, R; Rawlinson, L; Alder, J; (2005) Excavati ng sponges that are destructive 
to farmed pearl oysters in West ern and Northern Australia.  Aquaculture Research  36: 
150-162. 
Godwin, L S; (2003) Hull fouling of maritime vessels as a pathway for marine species 
invasions to the Hawai’ian Islands. Biofouling  19(1) Supplement 1: 123-131.  
Hein, F J; Risk, M J; (1975) Bioerosion of coral heads: inner patch reefs,  Florida reef tract.  
Bulletin of Marine Science  25: 133-138. 
Inglis, G J; Floerl, O; Unwi n, M; Ponder-Sutton, A; Seawar d, K; Kospartov, M; Bell, A; 
Kluza, D (in prep.). The biosecurity risks associated with biofouling on international 
vessels arriving in New Zealand. Summary of the patterns and predictors of fouling.  
Biosecurity New Zealand Technical Paper. 
Kospartov K; Inglis G; Seawar d K; Brink A van den; D’Arch ino R; Ahyong S (in press) Non-
indigenous and cryptogenic mari ne species in New Zealand . Current state of knowledge 
– Interim report prepared for MAFBNZ Post Border Directorate. MAF Biosecurity New 
Zealand. Wellington, New Zealand. 
Longo, C; Mastrototaro, F; Corriero, G (2007) Occurrence of Paraleucilla magna (Porifera: 
Calcarea) in the Mediterranean Sea.  Journal of the Marine Biological Association of 
the United Kingdom  87: 1749-1755. 
Probert, P K; Batham, E J (1979) Epibenthic m acrofauna off south-eastern New Zealand and 
mid-shelf bryozoan dominance.  New Zealand Journal of Marine and Freshwater 
Research  13: 379-392. 
Rowden, A A; O’Shea, S; Clark, M R (2002) Benthic biodiversity of seamounts on the 
northwest Chatham Rise .  Marine Biodiversity Biosecurity Report 2, Ministry of 
Fisheries; Wellington, New Zealand. 
Rützler, K; Muzik, K (1993) Terpios hoshinaota , a new cyanobacteriosponge threatening 
Pacific reefs. Scientia Marina  57(4): 395-403. 
Streftaris, N; Zenetos, A; Papathanassiou, E (2005) Globalisation in marine ecosystems: the 
story of non-indigenous marine sp ecies across European seas.  Oceanography and 
Marine Biology: An Annual Review  43: 419-453.