Use of Changes in Fish Assemblages in the Fly River System, Papua New Guinea, to Assess Effects of the Ok Tedi Copper Mine

Chapter 12 Use of Changes in Fish Assemblages in the Fly River System, Papua New Guinea, to Assess Effects of the Ok Tedi Copper Mine

Andrew W. Storey1,Ã, Markson Yarrao2, Charles Tenakanai3, Boga Figa3 and Jessica Lynas1

1School of Animal Biology (M092), The University of Western Australia, Crawley, WA, 6009, Australia 2Environment Department, Ok Tedi Mining Limited, PO Box 1, Tabubil, W.P. Papua New Guinea 3Livelihood Programs Department, Ok Tedi Mining Limited, PO Box 1, Tabubil, W.P. Papua New Guinea

12.1. Introduction

The Fly River in Papua New Guinea is one of the largest rivers in Australasia (mean annual discharge B6,000 m3 s1). It has a catchment area of 76,000 km2 and flows a distance of over 1,200 km from its source in the central highlands of New Guinea to the Gulf of Papua. Much of the catchment, particularly in the upper reaches, consists of dense primary tropical rainforest, while in the middle and lower reaches open savannah forest, swamp forest, and seasonally inundated grasslands predominate. Although the upper catchment extends to altitudes of more than 3,500 m, the majority of the drainage basin is low-lying and flat, to the extent that the port of Kiunga, which is 800 river km from the coast, is only 20 m above sea level. The combination of low topography and high rainfall has resulted in a broad floodplain with extensive shallow lake systems, occupying an area of 4.5 million ha, making it the largest wetland system in the country. The wetlands

ÃCorresponding author. Tel.: (618) 6488 1482; Fax: (618) 6488 1029; E-mail: [email protected] (A.W. Storey). 428 A. W. Storey et al. of the Fly River are highly productive and play a vital role in the ecology of the river system. The area is sparsely populated, with an average human density of one–two persons per square kilometer. Prior to 1960, the highlands area had little contact with the outside world, and the river downstream was relatively pristine, with no mining or logging, and only low-scale commercial fishing for barramundi. Following the discovery of the Mt Fubilan ore deposit, scientists began surveying the aquatic fauna of this previously unstudied, remote river system (Boyden et al., 1978; Roberts, 1978; DPI, 1979, 1980; Robertson and Baidam, 1983). These studies revealed that the river system has the most diverse freshwater fish fauna in the Australasian region, and is known to support over 115 freshwater and marine vagrant species (Roberts, 1978; Maunsell and Partners, 1982; Allen, 1991; Coates, 1993; Swales et al., 1999). Of these, 17 are endemic to the Fly Basin, and over 30 are known only from the Fly and one or more of the large rivers in central-southern New Guinea (Roberts, 1978). The fishes of the Fly River basin are characterized by the large individual sizes of some species, an abundance of endemic fishes, and the presence of species that are poorly represented in other parts of the world, particularly the ariid and plotosid catfishes. In most other ways, the composition of the freshwater fish fauna is largely determined by its position in the Australasian zoogeographical zone (Roberts, 1978; Coates, 1993). Given the uniqueness of the fish fauna and concerns over the potential impact of a mine on its headwaters, environmental monitoring of the Fly River for the proposed mine commenced in 1981 (Maunsell and Partners, 1982). Monitoring began with an impact assessment which involved an expeditionary survey of water quality, fish communities in different habitats and metal levels in biota (Maunsell and Partners, 1982). Soon after, Ok Tedi Mining Ltd (OTML) implemented an extensive biological monitoring regime that commenced during the mine construction phase in 1983 and included all river reaches, from the headwaters to the delta and into the Gulf of Papua (Wood et al., 1995; Swales et al., 1999). An important aspect of this program was monitoring fish populations, partly in recognition of their value as a tool for assessing anthropogenic impacts, and as an indicator of ecosystem health and productivity (Fausch et al., 1990; Harris, 1995). It is also because the fish fauna forms an important subsistence food source for village communities along the river (Hortle, 1986). Due to this reliance on fish, the main concern of the State of Papua New Guinea was that this resource had to be protected (OTML, 1988, 1990). As a result, monitoring tended to concentrate on numbers and weight of fish available at locations throughout the river system. This emphasis has continued through mine life, and has influenced sampling methods, data collected, and how analyzed. Fish Assemblages in the Fly River System and Effects of the Ok Tedi Copper Mine 429

There have been three operating stages in mine life to date. Stage One (the treatment of gold ores by cyanide extraction) commenced in 1984. Stage Two commenced in mid-1987 and consisted of the gold extraction circuit running in parallel with a flotation circuit to produce copper concentrate. Stage Three commenced in mid-1988 when the gold circuit was decommissioned (and all use of cyanide ceased) and the mine became a copper concentrate producer. Additional infrastructure was completed in 1989, enabling the mine to treat 80,000 t of ore per day at peak production. The next major change was in 1998 when a dredge was deployed on the lower Ok Tedi to dredge the channel with the intention of alleviating bed aggradation in the lower Ok Tedi and Middle Fly. Currently, approximately 50 Mt per annum (pa) of waste rock and 30 Mt pa of tailings are discharged directly into the Ok Tedi and its tributaries, and approximately 15 Mt pa of sediment are dredged from the lower Ok Tedi and deposited in ‘‘cells’’ on the adjacent floodplain. The original mine plan was to operate with riverine disposal of waste rock, but storage of tailings in a dam in the upper catchment. However, the tailings dam collapsed during construction, and to allow the mine to proceed, the State permitted riverine disposal of detoxified tailings. The Ok Tedi mine subsequently received notoriety in the mid-1980s when the cyanide detoxification system failed on several occasions, releasing tailings with elevated cyanide levels into the Ok Tedi, resulting in the death of fish, turtles, and crocodiles for approximately 100 km downstream. Since the closure of the gold extraction circuit (1988), the mine has operated with a copper flotation circuit, which recycles the flotation chemicals. As such, the tailings have no elevated chemicals, but are elevated in a suite of particulate and to a lesser extent dissolved metals including Cu, Pb, Zn, Cd, As, and Fe, compared to average crustal abundance (Bolton et al., 2009). Monitoring the effects of the Ok Tedi mine on the river system has continued to present, principally to document and understand impacts resulting from the mode of operation of the mine (Pickup and Cui, 2009). Relationships between changes in fish stocks in the main river channel and mine waste discharges have been reported (Smith et al., 1990; Smith and Hortle, 1991; Smith and Morris, 1992), as have impacts to the fish fauna of riverine and floodplain habitats (Swales et al., 1998, 1999, 2000). Since these publications, additional data have been collected and further analyses conducted. This chapter describes temporal variations in fish catch and assemblage composition, and updates previous assessments of species diversity and biomass from sites downstream of the mine (OTML, 1994, 1995, 1996; Swales et al., 1998, 1999, 2000). 430 A. W. Storey et al.

12.2. Sampling and Analyses

Although monitoring only commenced during mine construction, providing limited baseline data, the sampling method adopted at that time has been maintained throughout the project, providing a standard sampling approach, now covering almost 25 years. This provides an immensely valuable time series, unlikely to be matched from any other river in the world, let alone a large, tropical river system. Such intensive sampling is known to yield a high number of species and more accurate and precise estimates of richness (Cao et al., 2001; Hughes et al., 2002; Reynolds et al., 2003; Kennard et al., 2006). Methods were based on those used by Maunsell and Partners (1982) using a standard set of 13 gill-nets, ranging in stretched mesh size from 25 to 175 mm (Table 12.1). The selection of gill nets as the main method was a compromise over what was readily available and could be used and maintained in a remote region, and what was effective in the conditions and across habitats. It is also a sound, simple, and technically robust method to obtain catch per unit effort (CPUE) data, ideally suited to a remote location where failure of more technical equipment (i.e., electroﬁshers or hydro- acoustic gear) can result in many months delay. Seine netting, rotenone, trapping, and electroﬁshing were also used selectively in certain habitats to supplement catches; however, these data are not included in the current

Table 12.1: Gill nets used in the Standard Gill Net Set used by OTML since 1983, giving stretched-mesh size of each net, line type, number set, dimensions, and area.

Stretched-mesh size and type Number in Length (m) Depth Area net set (m) (m2) 1v (25 mm) Monofilament 1 40 2.3 92 1½v (38 mm) Monofilament 1 40 1.7 68 2v (50 mm) Monofilament 1 45 2.1 95 2½v (63 mm) Monofilament 1 40 2.8 112 3v (75 mm) Monofilament 1 45 3.2 144 3½v (88 mm) Monofilament 1 45 3.5 158 4v (100 mm) Monofilament 1 45 4.2 189 5v (125 mm) Monofilament 1 45 4.9 221 6v (150 mm) Monofilament (6M) 1 50 6 300 6v (150 mm) Multifilament (6C) 2 25 2.8 70 7v (175 mm) Multifilament (7C) 2 25 3.1 78 Fish Assemblages in the Fly River System and Effects of the Ok Tedi Copper Mine 431 analyses as they can confound CPUE data across sites and times. Monitoring was infrequent during the early years (1983–1988), but increased to quarterly during the late 1980s and 1990s, with monthly sampling at some locations. However, this frequency has since been reduced to biannually, and at fewer sites (2000 onward). Sites were initially selected to provide a good geographical spread and ease of access to satisfy monitoring programs designed to assess the state of the indigenous subsistence fishery, rather than to provide ecological data. Often sites were adjacent to villages with airstrips, which provided access to the river. Details of the sites sampled and their locations are presented in Table 12.2 and Fig. 12.1, respectively. Sites were stratified between riverine and floodplain habitats, respectively. Riverine sites ranged from fast-flowing, forested, cobble/gravel bed channel approx 100 m across and 2–3 m deep in the upper catchment, to slower flowing, meandering forested channel, 150 m wide and 5–10 m deep with silt/ sand substrate in the middle reaches, to slow-flowing sections meandering through seasonally inundated grassed floodplain, approximately 200–300 m wide, 10–20 m deep with fine silt/mud substrate. Floodplain sites were separated between forested floodplain in the upper half of the catchment and grassed floodplain in the lower half. Within these broad regions, there were sites in oxbow lakes, being cut-off meander bends of the river, as wide

Table 12.2: Key long-term monitoring sites and their period of sampling.

Site name Site code Start of sampling End of sampling Riverine sites Ningerum TED20 April 1983 March 1996 Atkamba TED30 December 1983 October 1992 New Atkamba TED35 January 1993 March 2002 Kuambit FLY10 June 1983 October 2003 Bosset FLY14 June 1983 February 2004 Obo FLY15 April 1987 February 2002 Ogwa FLY20 April 1987 February 2004 Strickland River STR01 April 1987 February 2002 Floodplain sites Bosset Lagoon BOS10 June 1983 May 1996 Daviumbu DAV01 April 1987 February 2002 Sembe Oxbow OXB03 October 1990 October 1998 Lake Pangua OXB05 June 1989 February 2002 Oxbow at ARM345 OXB06 June 1993 March 2002 Strickland OXB08 March 1992 February 2002 432 A. W. Storey et al.

Figure 12.1: Location of riverine and ﬂoodplain sampling sites on the Fly River system.

(150–300 m) and deep (10–20 m) as the adjacent river and lined with either primary rainforest or flooded grasses (Saccharum and Phragmites). The other main floodplain sites were blocked valley lakes, being broad, shallow lagoons 3–4 m deep and heavily vegetated around the margins with floating grasses, and submerged and emergent macrophytes. The blocked valley lakes recede Fish Assemblages in the Fly River System and Effects of the Ok Tedi Copper Mine 433 each dry season and dry totally in El Ninõ droughts, whereas the deeper oxbow lakes never dry out, providing drought refuge. At riverine sites, nets were set at approximately 301 to the bank in decreasing order of mesh size with distance upstream, and in floodplain sites they were set perpendicular to the bank, with large mesh nets at either end of the site, and small mesh nets set progressively toward the center. Nets were set at least 50 m apart. Nets of the same mesh size tended to be set at the same location at each site on each occasion (i.e., the same mesh size would be deployed at a specific backwater at a site on each sampling occasion) to minimize temporal variability. At each site, on each sampling occasion, the nets were deployed for 24 h, checked at dusk and dawn, and all fish caught identified to species. Fish length and weight were also recorded. This sampling method provided abundance, biomass, and assemblage composition data for analysis. Monitoring data were then analyzed to determine temporal changes in species richness, biomass, and assemblage composition. To assist detection of mine effects, and acknowledging the different phases of mining, data for each site were subdivided into time periods corresponding to different ‘‘mine operating periods’’ (Pickup and Cui, 2009) (Table 12.3). Temporal changes in species richness and biomass were examined using linear regression for each site. Biomass data were transformed (ln(xþ1)) prior to analysis. Where a significant relationship (po0.05) was found, percentage change was determined using the observed data and predictions from regression analysis (i.e., change from beginning of sampling period versus end of period). Changes in the frequency of occurrence of each species in each time period were tabulated and between-time period differences in observed over expected frequency of occurrence tested by Chi-square contingency table analysis (Zar, 1974). Multivariate analyses were also performed to detect changes in fish assemblages at sites over time. Ordinations were performed using the semi- strong hybrid multidimensional scaling (SSH MDS) procedure in the PATN pattern analysis software (Belbin, 1995), using abundance of each species at a site on each sampling occasion. Dissimilarity between samples at a site over times was determined using the Bray-Curtis dissimilarity coefficient, and species occurring in less than 10% of samples in any data set were omitted to avoid ‘‘rare’’ species having a disproportionate effect on the analyses (Gauch, 1982). Sites were analyzed separately, and samples were grouped according to time periods and mining method (Table 12.3) and these groups illustrated on the ordinations. Finally, principal axis correlation (PCC) (Belbin, 1995) was used to test for significant gradients in community descriptors (viz. species richness and biomass) through ordinations of each 3 .W trye al. et Storey W. A. 434 Table 12.3: Mine operating periods used to assess temporal changes in fish assemblage structure at key monitoring sites, indicating the number of samples in each period from each site.

Time periods

Baseline & gold Gold/copper & CopperW80,000 tpd CopperW80,000 tpd Dredge in operation only (Au) coppero80,000 (1–5 years) (6–10 years) (Dredge) tpd (Au & Cu) (CuW80k 5 year) (CuW80k 10 year)

Start of period December 1983 August 1987 August 1989 September 1993 March 1998 End of period July 1987 July 1989 August 1993 February 1998 March 2004 Site Riverine (code) Ningerum (TED20) 15 6 4 4 Atkamba (TED30) 13 9 23 New Atkamba (TED35) 8 30 35 Kuambit (FLY10) 14 9 43 49 33 Bosset (FLY14) 14 7 14 14 16 Obo (FLY15) 8 15 16 14 Ogwa (FLY20) 9 15 16 16 Strickland R. (STR01) 9 14 16 14 Floodplain (code) Bosset Lagoon (BOS10) 14 7 15 10 9 Daviumbu (DAV01) 7 12 12 10 Sembe Oxbow (OXB03) 12 12 4 Lake Pangua (OXB05) 14 14 14 Oxbow at ARM345 7813 (OXB06) Strickland (OXB08) 5 9 4

Note:Au¼ gold cyanide leachate phase; Au & Cu ¼ gold cyanide leachate and copper flotation circuit; CuW80k 5 year ¼ copper flotation circuit in isolation, running atW80,000 tpd for the first five years of this phase (1989–1993); CuW80k 10 year ¼ the second five years of this phase (1993–1998); and, Dredge ¼ copper flotation circuit with the dredge operating in the lower Ok Tedi. Fish Assemblages in the Fly River System and Effects of the Ok Tedi Copper Mine 435 site. The technique identifies changes in species richness and biomass across ordination space, with the gradient indicating the direction of increasing values of the parameter (i.e., increasing in the direction of samples with high richness/biomass). Gradients of time in days from start of sampling (i.e., as an indication of cumulative exposure to mine effects) were also tested. Although a comprehensive geochemical monitoring program is conducted by OTML, unfortunately the sites and sampling times do not match the fish sampling sites/times, and as a result there are no quantitative physicochem- ical data that exactly match individual fish catch data at each site on each occasion to assist interpretation. However, changes in fish assemblages correspond with gross patterns of changes in water quality reported elsewhere in this volume (Apte, 2009; Bolton et al., 2009; Pickup and Cui, 2009).

12.3. Changes in Biodiversity

A number of methods have been developed worldwide to evaluate anthropogenic disturbance in river systems, including assessments of macroinvertebrates, chemical variables, and fish diversity (Karr, 1991; Hugueny et al., 1996). While historically, assessments of chemical variables (i.e., nutrients, heavy metals, and salinity) were popular, they do not integrate environmental quality over time. More recently, fish diversity has been used as an indicator of ecosystem health worldwide (Karr, 1981; Oberdorff and Hughes, 1992; Hugueny et al., 1996; An and Choi, 2003). Because fish continually inhabit the receiving water, they integrate the chemical, physical, and biological histories of the river. Most fish species have a long life span and therefore reflect both long-term and current water quality. Sampling fish assemblages can be used to assess a range of environmental disturbances, such as changes in habitat, water quality, and land use (Hugueny et al., 1996). Fish are also one of the most conspicuous biota in the Fly River system, they are relatively easy to sample and identify, but they are also a significant indigenous resource.

12.3.1. Riverine Sites

Since monitoring began in 1983, 86 species representing 32 families have been recorded from sites on the Fly, Ok Tedi, and Strickland rivers (Swales et al., 2000). Of these, catﬁsh of the families Ariidae (16 spp) and Plotosidae 436 A. W. Storey et al.

(9 spp) were highly dominant in terms of diversity, but with Nematolosa herrings being the most abundant species (forming over 37% of the catch). A number of species known from the Fly River are considered important owing to their large size, including anchovy (Thryssa scratchleyi), catfish (Arius augustus), Papuan bass, (Lutjanus goldiei), and barramundi (Lates calcarifer). Barramundi provides valuable commercial and artisanal fisheries. Of importance, from both an ecological and social perspective, is the fact that declines in species diversity have been recorded from the majority of sites, with the exception of New Atkamba (TED35) and Strickland River (STR01, a control site) (Table 12.4 and Fig. 12.2). Since the commencement of monitoring, reductions in species diversity ranging from 21% (FLY20) to 80% (TED20) have been recorded (Table 12.4). The greatest reductions in the number of species were reported from the more heavily impacted sites, closest to the mine, i.e., Ningerum TED20 (80%) and TED30 (69%). Nonetheless, considerable declines in the number of fish species collected were also reported from the middle Fly at Bosset FLY14 (46%) and Obo FLY15 (35%). Sampling at TED35 commenced in January 1993, as a replacement for Atkamba (TED30), which was effectively lost when the meander loop was cut off as an oxbow lake. It is likely that the short sampling period for TED35 limited the ability to detect mine effects, but particularly since the period of record covers only the copper flotation period, and the lower Ok Tedi had already been heavily impacted prior to this time. Future monitoring of this site is important since it is the closest riverine site downstream of the dredge. The dredge was implemented to remove sediment from the lower Ok Tedi, prior to discharge into the Fly River and thereby reduce bed aggradation and alleviate overbank flooding and associated environmental impacts downstream. Chi-square contingency table analysis of changes in the frequency of occurrence of individual species from each site over time, showed that the greatest loss and declines in occurrence of species occurred at sites closest to the mine (i.e., TED30, FLY10, FLY14, and FLY15) (Table 12.5). Species particularly noted for their declines from riverine sites include many of the forktailed catfishes; Arius latirostris, Arius carinatus, Arius augustus, Cinetodus crassilabris, Arius macrorhynchus, Cinetodus froggatti, and Nedystoma dayi. The mullet Liza alata (diadema) has also declined at the most impacted riverine sites. Most of these species are resident in the river system, except the mullet, which likely has an estuarine/marine affinity. Little is known of the biology of the individual species, which limits the ability to interpret reasons for their declines; however, issues such as increased suspended sediment loads, mobile bed loads, smothering of habitats, loss of ihAsmlgsi h l ie ytmadEfcso h kTd oprMn 437 Mine Copper Tedi Ok the of Effects and System River Fly the in Assemblages Fish

Table 12.4: Regression statistics and percent reduction in species richness and biomass from riverine and ﬂoodplain sites.

Site Reduction in species richness Reduction in biomass

pr-square % pr-square % reduction reduction

Riverine Ningerum (TED 20) po0.0001 0.620 90 po0.005 0.257 86 Atkamba (TED30) po0.0001 0.485 75 po0.0001 0.515 87 New Atkamba (TED35) pW0.05 – None pW0.05 – None Kuambit (FLY10) po0.0110 0.043 22 po0.0001 0.272 79 Bosset (FLY14) po0.0001 0.463 46 po0.0001 0.278 75 Obo (FLY15) po0.0001 0.262 35 p ¼ 0.0008 0.200 57 Ogwa (FLY20) p ¼ 0.0048 0.133 21 pW0.05 – None Strickland R (STR01) pW0.05 - None p ¼ 0.0075 0.132 49 Floodplain Bosset Lagoon (BOS10) p ¼ 0.0112 0.115 32 p ¼ 0.0021 0.165 72 Lake Daviumbu (DAV01) p ¼ 0.038 0.106 26 p ¼ 0.044 0.100 64 Sembe Oxbow (OXB03) pW0.05 – None pW0.05 – None Lake Pangua (OXB05) p ¼ 0.0157 0.137 25 p ¼ 0.0102 0.154 66 Oxbow ARM345 (OXB06) p ¼ 0.0082 0.240 25 p ¼ 0.0103 0.228 52 Strickland Oxbow (OXB08) pW0.05 – None pW0.05 – None 438 A. W. Storey et al.

Figure 12.2: Regressions of species richness against time at riverine sites, showing significance level, r-square, percent reduction from start of sampling, and 95% prediction intervals around the linear regression line. food resources, and possible issues such as chronic toxicity to eggs/larvae could all play a role in their decline. Concomitant with declines in fish biodiversity has been the deposition of mine-derived waste sediment, resulting in elevated sediment loads, increased particulate, and dissolved copper levels, die-back of floodplain forest and Table 12.5: Summary of changes in species occurrences at each riverine site. 439 Mine Copper Tedi Ok the of Effects and System River Fly the in Assemblages Fish

Family Species TED30 TED35 FLY10 FLY14 FLY15 FLY20 STR01

Ambassidae Ambassis agrammus ns ns m ns ns ns Ambassis macleayi –––ns––– Ambassis spp ––ns–––– Parambassis gulliveri k ns k ns ns ns ns Anabantidae Anabas testudineus –nsmmns ns ns Apagonidae Glossamia aprion ns ns - ns ns ns ns Glossamia trifasciata ––ns–––– Ariidae Arius augustus – k k ns Arius agreutes – ns – – – ns ns Arius berneyi ns m ns ns ns ns ns Arius carinatus ns kkns kk Arius latirostris k ns k ns ns Arius leptaspis ns m ns ns ns ns - Arius macrorhynchus kkkkk-ns Arius taylori ns - ns – – – ns Arius sp. A ––––––ns Cinetodus crassilabris ns ns k kk Cinetodus froggatti ns kk ns ns Cochlefelis spatula ns ns kkns ns ns Cochlefelis danielsi ns– nsnsnsnsns Nedystoma dayi – k ns ns ns Tetranesodon conorhynchus –––––nsns Atherinidae Craterocephalus randi –ns- –––– Belonidae Strongylura kreffti ns ns - ns - ns ns Carcharhinidae Carcharhinus leucas –––––ns– Centropomidae Lates calcarifer k ns k ns ns ns ns 4 .W trye al. et Storey W. A. 440 Table 12.5: (Continued ).

Family Species TED30 TED35 FLY10 FLY14 FLY15 FLY20 STR01

Clariidae Clarias batrachus –nsmmns ns ns Clupeidae Clupeoides papuensis ns ns ns ns ns – ns Nematalosa spp km-ns ns ns ns Datnioididae Datnioides quadrifasciatus ns ns ns ns ns ns ns Eleotridae Mogurnda cingulata ––ns–––– Mogurnda mogurnda – – ns – ns – – Ophieleotris aporos –––––ns– Oxyeleotris ﬁmbriata ns – ns ns ns ns Oxyeleotris herwerdini –nsmmns ns ns Oxyeleotris nullipora – – – – ns ns – Oxyeleotris lineolatus ns ns ns ns ns ns ns Oxyeleotris spp –––ns––– Engraulidae Thryssa rastrosa –nsk ns Thryssa scratchleyi ns - ns ns ns ns Thryssa spp –––––ns– Gobiidae Glossogobius concavifrons –ns –––– Glossogobius giurus –nsns–––– Glossogobius sp. ––ns–––– Hemiramphidae Zenarchopterus novaeguinae ns ns ns – – ns – Kurtidae Kurtus gulliveri – ns – ns ns Lutjanidae Lutjanus argentimaculatus ns – k ns ns Lutjanus goldiei –nsk ns ns ns ns Megalopidae Megalops cyprinoides ns ns ns kkns ns Melanotaeniidae Melanotaenia splendida ns ns - ns - ns – Mugilidae Crenimugil labiosus ns–nsns––– Liza alata (diadema) –nskkkns ns Liza macrolepis –––––ns– ihAsmlgsi h l ie ytmadEfcso h kTd oprMn 441 Mine Copper Tedi Ok the of Effects and System River Fly the in Assemblages Fish Liza subviridis – – ns ns – – – Osteoglossidae Scleropages jardini –nsk ns ns ns ns Plotosidae Neosilurus ater ns --ns ns ns ns Neosilurus equinus ns ns ns – – – – Neosilurus sp.C ns–ns–––– Oloplotosus luteus ns–nsns––– Plotosus papuensis – k ns ns – ns – Porochilus obbesi –nsm ns ns - ns Porochilus meraukensis ––k ––ns Porochilus spp –––ns––– Pristidae Pristis microdon –––ns ns Scatophagidae Scatophagus argus – – – nsnsns– Sciaenidae Nibea semifasciata ns ns k ns ns ns Sparidae Acanthopagrus berda ns – k ns ns ns ns Terapontidae Amniataba percoides –nsns ns ns ns Hephaestes fuliginosus – – ns ns – – – Hephaestes roemeri ns ns ns – – Pingalla lorentzi ns ns - ns – ns – Terapon lacustris –nsm ns ns ns ns Toxotidae Toxotes chatareus k ns - ns ns ns ns Toxotes lorentzi ––– ns ns ns No.of species lost ( ) 7238531 No. of species in decline (k) 62158432 No. of species increasing (m) 0363000 No. of species with sign 02100221 variability (-)

Notes: ¼species previously common, but not recorded in last time period; k ¼ species still present but showing a significant decline (po0.05) according to Chi-square test; - ¼ a significant change between time periods, but with lowest occurrences in intermediate years; ns ¼ no significant change in occurrence according to Chi-Square test; ‘‘–’’ ¼ species never recorded from site. NB analysis was not conducted for TED20 because this site was seldom sampled after 1987 because fish catch had declined so severely. 442 A. W. Storey et al. aggradation of the riverbed. Previous studies have ruled out acute copper toxicity to fish, as levels of dissolved copper were not sufficiently elevated to exceed riverine complexing capacity; labile copper levels were generally low (Smith et al., 1990; Smith and Hortle, 1991; Stauber, 1995). Habitat loss due to riverbed aggradation is likely one major reason for reductions in fish biodiversity from riverine sites of the Ok Tedi and Fly River (Storey et al., 2009). As sediment is transported and deposited downstream, diversity of potential fish habitat is reduced through the infilling of backwaters, smothering of woody debris, reduction in bank profile, and loss of emergent vegetation and channel heterogeneity. In addition, fish may be affected by the increased water velocity, and indirectly by changes to food web structure, including declines in invertebrate diversity, reduced terrestrial inputs (insects and fruits) from the forest, and likely loss of in-stream algal production (Storey and Yarrao, 2009). The changes in occurrence of some species may not be mine related. For example, Allen (1991) recorded the freshwater saw fish, Pristis microdon,as being common from the middle Fly River, but over the last 15 years this species has not been collected upstream of Everill Junction. It is, however, still commonly caught in gill nets downstream of this point and within the Strickland River. Owing to its elongate snout and rows of laterally projecting teeth, P. microdon is easily caught in gill nets. It is because of its morphology that it is likely to have been ‘‘fished-out’’ of the middle Fly by the increased incidence of gill netting by villagers. Because of its susceptibility to fishing pressure and habitat loss across its range, this species has been listed as ‘‘endangered’’ on the IUCN Redlist (2004). However, it is also equally plausible that this species is susceptible to mine impacts and avoids the Fly River upstream of the major point of dilution at Everill Junction. It is generally accepted that there has been a loss of species from the Ok Tedi (Wood et al., 1995). However, it was previously suggested that all species were either still present in tributaries or within the main river and therefore, the system retained its capacity to return to pre-mine composition (Storey, 1997). Current data and analyses indicate however that some species are becoming even more restricted in their distribution and may be missing from the Ok Tedi and the Middle Fly. This would seriously restrict the capacity of the system to recover post-mine closure. In response to this concern, an extensive fish diversity survey was conducted in mid-2005 (WRM and Hydrobiology, 2007). The study showed that numbers of migratory species were much reduced in the middle and upper Fly, including Barramundi (Lates calcarifer), Oxeye herring (Megalops cyprinoides), mullet species (Liza diadema/alata, Crenimugil heterocheilus), Croakers (Nibea squamosa), Pikey bream (Acanthopagrus berda), gobies (Glossogobius spp.), Fish Assemblages in the Fly River System and Effects of the Ok Tedi Copper Mine 443 snappers (Lutjanus goldei, L. argentimaculatus), Soles (Asseraggodes klunzingerii), and Tongue soles (Cynoglossus heterolepis). Many of the Ariidae (forktailed) catfish species expected to be taken from the middle Fly were absent or rarely caught, such as A. latirostris, Co. spatula, Co. danielsi, A. taylori (robertsi), A. augustus, Ci. carinatus, Ci. froggatti, Ci. crassilabris, and N. dayi. This probably reflects mine-related impacts. Although these catfish were absent/rare in the middle Fly, most of these species were found elsewhere in the river system, either in the lower Fly/Strickland or the upper Fly tributaries (Elevala/Palmer). However, fishing pressure, especially in the upper Fly may undermine the value of this area as a refuge area. Tributaries of the upper Fly (Palmer/Elevala) and the larger Ok Tedi tributaries (Ok Mart, Ok Menga) continued to support a relatively diverse range of species, and continue to provide a refuge area from which species could colonize the middle Fly/Ok Tedi post-mine closure. However, abundances in these tributaries were generally lower than occurred in the Middle Fly, increasing the risk of loss of genetic diversity. The Fly River below Everill Junction also supported species absent/rarely taken from other parts of the system, but this area is less secure, especially if mine-related impacts progress into the lower Fly (WRM and Hydrobiology, 2007).

12.3.2. Floodplain Sites

The Fly River floodplain has been found to support diverse and abundant assemblages of freshwater fish, with 66 species from 33 families being recorded (Swales et al., 1999). In a similar way to the main river channel, fish diversity on the floodplain was found to be dominated by catfish of the families Ariidae (11 species) and Plotosidae (7 species). Species richness ranged from 32 (OXB06) to 46 (BOS10). Generally, fish diversity was more consistent at oxbow sites in comparison to blocked valley lakes and grassed floodplain sites. Marked inter-annual variations were evident from the latter sites, with a notable decline in diversity recorded during 1993/94. This coincided with El Ninõ drought conditions which resulted in very low river levels, with most habitats on the floodplain drying. The exception was the oxbow lakes, which, being deeper (generally 10–20 m deep), provided a drought refuge (Smith and Bakowa, 1994). Observed differences in fish communities from the different floodplain habitats likely reflect differences in physical habitat, and the tendency for the grassed floodplain and shallower blocked valley lakes (3–4 m deep) to dry out in times of drought. 444 A. W. Storey et al.

Since monitoring began in 1983, reductions in fish biodiversity have also been detected from floodplain sites. In fact, at four of the seven sampling sites, a significant decline in species diversity has been recorded since the commencement of biological monitoring. However, reductions were generally less severe than was noted for riverine sites, and ranged from 25% (OXB05 and OXB06) to 32% (BOS10) (Table 12.4 and Fig. 12.3). The site closest to the Ok Tedi (OXB06) experienced a 25% reduction in species diversity. Chi-square contingency table analysis showed that within the floodplain sites, the highest rates of loss and decline in species were from Lake Pangua (OXB05), Bosset Lagoon (BOS10), Sembe Oxbow (OXB03), and Oxbow at ARM345 (OXB06) (Table 12.6). Species which recorded consistent declines at floodplain sites included Oxyeleotris fimbriata and Oxyeleotris lineolatus.

Figure 12.3: Regression of species richness against time for floodplain sites, showing significance level, r-square, percent reduction from start of sampling, and 95% prediction intervals around the linear regression line. Table 12.6: Summary of changes in species occurrences at each floodplain site.

Family Species BOS10 DAV01 OXB03 OXB05 OXB06 OXB08 445 Mine Copper Tedi Ok the of Effects and System River Fly the in Assemblages Fish

Anabantidae Anabas testudineus m ns ns ns ns ns Ambassidae Ambassis agrammus ns ns ns ns ns ns Ambassis macleayi ns – – ns ns Ambassis spp –ns––ns– Denariusa bandata nsns–––– Parambassis gulliveri ns ns ns ns ns ns Apagonidae Glossamia aprion - ns ns ns ns ns Glossamia trifasciata –ns–––– Ariidae Arius agreutes –––––– Arius augustus ns – – ns – ns Arius berneyi ns - ns ns ns ns Arius carinatus –ns–––ns Arius latirostris –––––– Arius leptaspis ns ns ns ns ns ns Arius macrorhynchus –––––– Arius taylori –––––– Arius sp.A –––––– Cochlefelis spatula ns – – ns – – Cochlefelis danielsi –––ns–ns Cinetodus crassilabris –––––ns Cinetodus froggatti - –ns–ns Nedystoma dayi ––ns––m Tetranesodon conorhynchus –––––– Atherinidae Craterocephalus randi – – ns ns – – Belonidae Strongylura kreffti - ns ns ns ns ns Carcharhinidae Carcharhinus leucas –––––– Centropomidae Lates calcarifer k ns k-ns ns Clariidae Clarias batrachus ns ns ns ns – – Table 12.6: (Continued ). 4 .W trye al. et Storey W. A. 446 Family Species BOS10 DAV01 OXB03 OXB05 OXB06 OXB08 Clupeidae Clupeoides papuensis –––––– Nematalosa spp - ns ns ns ns ns Datnioididae Datnioides quadrifasciatus ns ns ns k ns Eleotridae Mogurnda mogurnda –––––– Mogurnda cingulata –––––– Ophieleotris aporos –ns–––– Oxyeleotris ﬁmbriata ns ns Oxyeleotris nullipora –––––– Oxyeleotris lineolatus ns ns ns Oxyeleotris herwerdini mmmmns ns Oxyeleotris spp –––––– Engraulidae Thryssa scratchleyi k--ns ns ns Thryssa rastrosa Thryssa spp ns – ns – ns ns Gobiidae Glossogobius giurus –––––– Glossogobius sp. –––––– Glossogobius concavifrons –––––– Stenogobius lachneri –––ns–– Hemiramphidae Zenarchopterus novaeguinae –––ns–– Kurtidae Kurtus gulliveri ns – ns – ns Lutjanidae Lutjanus argentimaculatus ns ns – – ns ns Lutjanus goldiei ns ns ns k ns ns Melanotaeniidae Melanotaenia maccullochi ––ns––– Melanotaenia splendida ns ns ns ns ns ns Megalopidae Megalops cyprinoides ns ns ns k ns ns Mugilidae Crenimugil labiosus –––––– Liza alata (diadema) ns - k ns ns Liza macrolepis –––––– Liza subviridis –––––– Osteoglossidae Scleropages jardini k ns ns ns ns ns Plotosidae Neosilurus ater ns ns ns ns - ns Neosilurus equinus –––––– 447 Mine Copper Tedi Ok the of Effects and System River Fly the in Assemblages Fish Neosilurus sp.C –ns–––– Neosilurus brevidorsalis ns ns ns ns – – Plotosus papuensis ns ns – – ns ns Oloplotosus luteus –––––– Porochilus obbesi ns ns ns - ns Porochilus meraukensis m ns ns ns k Porochilus spp – ns ns – ns – Pristidae Pristis microdon ––––ns Scatophagidae Scatophagus argus –––ns–– Sciaenidae Nibea semifasciata –––nsnsns Sparidae Acanthopagrus berda –ns–k –ns Terapontidae Hephaestes roemeri –––––– Hephaestes fuliginosus –––––– Pingalla lorentzi ns ns – ns ns – Amniataba percoides --ns --ns Terapon lacustris ns ns m ns ns ns Terapon jarbua –ns–––– Toxotidae Toxotes chatareus ns -kns ns ns Toxotes lorentzi ns ns –– No. species lost ( ) 634532 No. species in decline (k) 302420 No. species increasing (m) 312101 No. species with sign 461230 variability (-)

However, since Oxyeleotris herwerdini increased in occurrence at most sites from which the former species declined, and these species are very similar, it is likely the reported declines are actually due to misidentification, although species replacement cannot be discounted. The mullet Liza alata (diadema) has also declined at a number of floodplain sites (i.e., OXB05, OXB03, and BOS10), probably reflecting the declines in riverine sites. Natural environmental change likely explains the significant declines in species diversity reported from Daviumbu (DAV01), Pangua (OXB05), and Bosset (BOS10). El Ninõ droughts in recent years (1983, 1986, 1992, 1993, and 1997), caused large areas of the floodplain to dry out with adverse effects on fish communities through habitat loss and associated changes in water quality. These effects can be seen in the monitoring data. Chi-square analysis showed that for a number of species, frequency of occurrence reduced in intermediate years, but subsequently recovered (i.e., Arius berneyi, Porochilus obbesi, Porochilus meraukensis, and Amniataba percoides). This undoubtedly reflects declines in populations of these species at blocked-valley lake sites in El Ninõ years, with subsequent lags in recovery. Lakes Daviumbu and Bosset, as with other blocked valley lakes dry during El Ninõ droughts. As they re-flood they become covered in a continuous, dense (1 m thick) mat of floating grass (wild rice – Oryza sp.) which seems to prevent recolonization by fish. This is likely due to low light and low dissolved oxygen levels under the grasses. Assuming the lagoon does not dry in the intervening time, it appears to take 12–18 months for the floating grasses to break up, disperse, and die off. It is not until the water body is predominantly open water that fish numbers and diversity start to recover. These changes do not appear at oxbow lake sites, because they are much deeper and never dry. However, changes observed at Lake Pangua appear to be due to a different process. This oxbow is very deep (þ22 m) and becomes strongly stratified in oxygen and temperature (AWS, unpublished data). The observed algal blooms and declines in fish catch are likely the result of the occasional breakdown of stratification and mixing of the water column leading to algal blooms and intermittent anoxia. This process appears to recur in Lake Pangua, but not in other oxbow lakes. The exact reason is not known, but could be because it is deeper than most oxbow lakes (W22 m), which may result in stronger stratification and anoxia of hypolimnetic waters; there may also be nutrient enrichment at this oxbow from a nearby village, which helps drive algal blooms. While in earlier years all changes in fish structure of floodplain sites were attributed to natural events and climatic phenomena (Swales et al., 1999), mine-related declines in fish diversity are now becoming apparent (viz. OXB06). A number of explanations associated with mine operations have Fish Assemblages in the Fly River System and Effects of the Ok Tedi Copper Mine 449 been postulated. As a result of bed aggradation in the main channel, there is increased frequency and duration of inflow to the oxbow lakes. This results in increased suspended sediment (TSS) levels in the water column, increased dissolved metal levels with the potential for chronic or possibly acute toxicity to fauna, and loss of riparian vegetation around the water bodies associated with forest die-back. A number of investigations are currently underway to try to separate the possible effects of these stressors. Riverbed aggradation likely impacts floodplain habitats through overbank flooding, inputs of mine-derived sediments, and die-back of flood-intolerant vegetation (Pickup and Cui, 2009). Recent analysis has detected a change in the algal contribution to the food web of OXB06, which may be related to die-back and sediment inputs (Storey and Yarrao, 2009). An issue likely to affect floodplain (and riverine) habitats is the development of localized acid rock drainage (ARD) on levee banks and tie-channel banks of the middle Fly (Bolton et al., 2009; Apte, 2009). Oxidization (exposure to air) of acid- generating sediments on levees and subsequent wetting and leaching lead to the release of metals, such as copper, into adjacent surface waters. This process seems particularly prevalent at the end of a dry period, with the onset of localized rainfall which ‘‘leaches’’ the levees. This has resulted in highly elevated levels of dissolved metals in surface waters immediately adjacent to the levees. Tie-channels connecting oxbow lakes to the main channel are particularly susceptible to deposition of mine-derived sediments and the establishment of ARD. It is also known that fish are sensitive to increases in heavy metals (including copper) and show an active preference for waters with low metal levels (Sorrenson, 1991). Therefore, fish may avoid areas leaching metals, including tie-channels, dissuading migratory species such as mullet, barramundi, ox-eye herring, and bass from entering these water bodies. Reductions in species diversity within ORWBs is of particular concern since it suggests fish lost from riverine sites are not using floodplain habitats as refuges against adverse conditions in the Ok Tedi and Fly rivers. This perhaps implies the loss of species from parts of the Fly River system, with implications for the maintenance of a subsistence fishery reliant on these species, and also for ecosystem recovery following mine closure.

12.4. Changes in Fish Catch (Biomass)

At the onset of mine development, the importance of the ﬁshery resource of both the river channel and ﬂoodplain to local people was recognized. In fact, 450 A. W. Storey et al.

fish catch was set as a compliance indicator, to be reported to the State of PNG on a regular basis to ensure protection of the subsistence fishery (OTML, 1988, 1990). Emphasis has therefore been placed on analyzing monitoring data to assess changes in fish catch over time, using total biomass of fish taken in the standard 24 h gill net set.

12.4.1. Riverine Sites

Over the entire monitoring period, considerable reductions in biomass have been detected from the majority of riverine sites. Significant reductions in biomass ranged from 49% at the Strickland River site (STR01) to 92% in the Ok Tedi (TED20) (Table 12.4 and Fig. 12.4). The site on the Ok Tedi in closest proximity to the mine (TED20) recorded the most extreme reduction in biomass (92%), closely followed by the next downstream site (TED30, 88%). Within the Fly River, the extent of declines in biomass decreased with increasing distance from the mine, i.e., Kuambit FLY10 (79% reduction), Bosset FLY14 (75%), and Obo FLY15 (57%) (Table 12.4 and Fig. 12.4). Barramundi, which formed a high proportion of catch biomass at many sites, particularly in the middle Fly, declined in number at most sites following peak numbers in the early 1990s following a period of good recruitment of this catadromous species from coastal areas. While Ogwa (FLY20) showed a reduction in species diversity over time, it did not show declines in biomass. This can perhaps be explained by the remaining fish at this site being comprised of large species. However, the linear regression plot suggests a trend in decreasing fish catch, with the last three sampling events recording lower than average biomass. If this trend continues, declines may become significant in the future. Interestingly, fish biomass also decreased at the site on the Strickland River over the monitoring period (STR01); however, this site exhibited the lowest level of decline. STR01 was originally selected as a control site, but the subsequent development upstream of the Porgera gold mine has meant that it can no longer be regarded in this way. The possible effects of the Porgera mine on the Strickland River are unknown.

12.4.2. Floodplain Sites

Since 1983, ﬁsh catch has generally been higher from oxbow lake sites (mean biomass range 214–280 kg) than at either blocked valley lakes (54–144 kg) or ﬂoodplain sites (47–104 kg). Such differences can perhaps be explained by the Fish Assemblages in the Fly River System and Effects of the Ok Tedi Copper Mine 451

Figure 12.4: Regressions of fish catch biomass against time at riverine sites, showing significance level, r-square, percent reduction from start of sampling, and 95% prediction intervals around the linear regression line. habitat at these sites, with the shallower, well-vegetated blocked valley lakes supporting more of the smaller fish species (i.e., Iriatherina werneri, Ambassis agrammus, and Melanotaenia sp), and the deeper oxbow lakes supporting a greater number of the larger predatory species, such as Lates calcarifer and Scleropages jardinii. In 1995–1996, 11 species recorded from the oxbow lakes 452 A. W. Storey et al.

(Cinetodus froggati, Nedystoma dayi, Lates calcarifer, Parambassis gulliveri, Clarias batrachus, Ophieleotris aporos, Thryssa rastrosa, Thryssa scratchleyi, Kurtus gulliveri, Lutjanus goldiei, and Pristis microdon) were not recorded from blocked valley lakes. Only one species (Iriatherina werneri) recorded from blocked valley lakes was not found in oxbow lakes. Some of these differences may be influenced by El Ninõ droughts, drying of wetlands, and rates of recolonization. A large number of floodplain sites also showed trends of decreasing biomass (Table 12.4 and Fig. 12.5). In fact, a significant decline in fish biomass was detected from most sites, with reductions as high as 72% (BOS10). Earlier suggestions of natural environmental change no longer account for all reported changes in fish catch, with sites in close proximity to the mine increasingly showing effects (i.e., OXB06). Possible natural and

Figure 12.5: Regressions of fish catch biomass against time at floodplain sites, showing significance level, r-square, percent reduction from start of sampling, and 95% prediction intervals around the linear regression line. Fish Assemblages in the Fly River System and Effects of the Ok Tedi Copper Mine 453 mine-related causes for declines in biomass are the same as those reported above for reductions in biodiversity in floodplain habitats.

12.5. Changes in Assemblage Composition

Much of the previous work on ﬁsh fauna of the Fly River and its ORWBs has utilized univariate approaches for analyses, with the exception of Smith and Morris (1992) and OTML (1993). Although univariate analyses provide an indication of how an individual species may change temporally or spatially, they do not indicate how ﬁsh assemblage structure as a unit may have changed. The total loss and subsequent replacement of a species (i.e., by an introduced species) would not be detected using univariate methods, for example, since the total number of species would remain unchanged. How- ever, multivariate analysis is able to identify a change in overall assemblage structure, statistically representing the degree of change (percentage similarity) between assemblages (Belbin, 1995). Multivariate approaches are widely used and accepted in community ecology (Gauch, 1982).

12.5.1. Riverine Sites

All sites showed significant separation among the different mine-operating time periods (Analysis of Similarity; ANOSIM, po0.05). Generally, samples from the most recent mine-periods (i.e., CuW80k & Dredge) displayed the greatest separation, with earlier time periods showing some overlap (Fig. 12.6). This suggests fish assemblage change has been more pronounced in recent times. Multidimensional Scaling (MDS) supported results from linear regression, with Ok Tedi (TED20 and TED30) and Fly River sites closer to the mine (FLY10, FLY14, and FLY15) showing reductions in species diversity and biomass over the monitoring period. This was evident from the significant gradients in time from these sites, indicating a progressive change in assemblage structure over time (Fig. 12.6). Further- more, gradients in biomass and species diversity were directly opposite to the time gradient, indicating a decline in these parameters over time (Fig. 12.6). Sites geographically further from the mine recorded less change in their fish assemblages than those from sites closer to the mine. Gradients in biomass and diversity from ordinations of the Fly River at Ogwa (FLY20) and the Strickland River (STR01), for example, were not directly opposite to the time gradient, indicating less severe changes to assemblage structure than reported from impacted sites (Fig. 12.6). 454 A. W. Storey et al.

Figure 12.6: MDS ordination of species assemblage data from riverine sites, indicating signiﬁcant gradients in assemblage descriptors and time.

New Atkamba (TED35), although on the Ok Tedi and close to the mine site, showed no significant change in either species richness or biomass using linear regression, as discussed above. Multivariate analyses did however detect a significant difference in fish assemblages between dredge and pre- dredge operating periods (ANOSIM, po0.05). Significant gradients in Fish Assemblages in the Fly River System and Effects of the Ok Tedi Copper Mine 455 species richness, biomass, and time were in the direction of samples collected post-dredge construction (Fig. 12.6). This infers an increase in species richness and biomass at this site following implementation of the dredge, suggesting possible improvements in fish assemblages.

12.5.2. Floodplain Sites

Floodplain sites generally demonstrated less change in fish assemblage structure than riverine sites. The majority of sites along the floodplain did not record any significant separation of mine-operating periods. Further- more, gradients in time were not significant for most sites, indicating temporal changes in assemblage composition were not great (Fig. 12.7).

Figure 12.7: MDS ordination of species assemblage data from ﬂoodplain sites, indicating signiﬁcant gradients in assemblage descriptors and time. 456 A. W. Storey et al.

The floodplain site closest to the mine (OXB06) appears to be showing signs of a developing change. Although separation of mine-operating periods was not significant (ANOSIM, pW0.05), samples collected between April 1999 and March 2002 (within the dredge period) showed considerable separation from the majority of samples in earlier time periods and from other samples in the dredge period (Fig. 12.7). The significant gradient in time was in the direction of this group of samples, indicating a recent, but consistent change in fish assemblages in this oxbow. Gradients in species richness and biomass were also significant, but were perpendicular to the time gradient, indicating a weak reduction in these parameters over time. Lake Pangua (OXB05) not only recorded a significant gradient in time, but also significant gradients of species richness and biomass which were in the opposite direction to time through the ordination (Fig. 12.7). This suggests that fish assemblages from Lake Pangua have declined in species richness and biomass over the monitoring period. Although changes to fish assemblages in this system are not entirely understood, and may be the result of environmental (El Ninõ, algal blooms, and anoxia) and social conditions (artisanal fishing and input of nutrients from the nearby village), the possibility of a mine-related impact should not be discounted. The oxbow at Sembe (OXB03) did have a significant gradient in time, indicating a systematic change over time. Gradients in biomass were also significant, and the earliest samples (CuW80K, 5 years) tended to separate from samples collected during later time periods (Fig. 12.7). As with Lake Pangua, this may also be mine related, but more data are required.

12.6. Conclusions

Analyses presented here, combined with results from a recent fish diversity study of the Fly River system (WRM and Hydrobiology, 2007) demonstrated that the majority of fishes of the Fly River system continue to maintain populations within the catchment, but there has been a marked reduction in the diversity and biomass of fishes in most reaches downstream of the Ok Tedi mine, as far as Everill Junction. Mine closure is currently planned for 2013; however, because of the volume of waste material stored in the upper catchment, and the rate of erosion and sediment transport through the system, the full impacts in the middle and lower Fly River will not be apparent for up to 50 years after mine closure (Pickup and Cui, 2009). It is therefore likely that impacts in the middle Fly may become worse, and fish populations downstream of Everill Junction may be affected in the future. Fish Assemblages in the Fly River System and Effects of the Ok Tedi Copper Mine 457

More than half the fish species formerly resident in the habitats of the Ok Tedi and Middle Fly sub-catchments are effectively no longer resident there. Populations of most species still occur in other parts of the catchment, and this provides the potential for recovery of the fish assemblages post-mining. However, four species (Clupeoides venulosus, Glossogobius celebius, Oxyur- ichthys papuensis, and Hephaestus raymondi) have not been found in the Fly River system since 2000 (WRM and Hydrobiology, 2007). If lost, this obviously limits the potential of the system to recover post-mine closure. The impacts of the mine on the fishery are indisputable; however, an additional factor that will further limit the ability of the system to recover post-mine is the presence of three exotic fishes – Climbing perch (Anabas testudineus), Striped snakehead (Channa striata), and Walking catfish (Clarias batrachus). These species have all invaded the Fly River system since 1988 (Storey et al., 2002). They were most likely brought from Indonesia by transmigrants settling in West Papua close to the border with Papua New Guinea. From here they likely escaped into the Trans-Fly floodplain system. All three are now well established throughout the Fly River, from the Lower Fly to the upland streams along the Kiunga-Tabubil Highway. This is of concern with respect to the potential recovery of native fish populations after mine closure. Exotic fishes commonly have competitive advantages over native fish species in river systems that have been highly modified (see discussions in Storey et al., 2002 and Dudgeon and Smith, 2006). The Fly River is now a highly modified system, with substantial habitat alteration resulting from the operations of the Ok Tedi mine, but also from the increased human population attracted by the mine. There are many new settlements downstream of the mine, with associated pressures on water courses as a result of clearing for gardens, everyday activities (e.g., washing, disposal of effluent, etc.), and increased fishing pressure. It can be expected that these exotic species will exert adverse pressure on the recovery of native fish populations. A recent study on the Fly River (August 2007) noted these exotic species in creeks and backwaters, and where exotics were prevalent native species were almost absent (A.W. Storey, personal observation). Unfortunately, it will not be possible to eradicate these exotic species, and the long-term impacts of these (and potentially other) introduced species on the native fish fauna of the Fly River will likely be far greater than the relatively short-term impacts of the Ok Tedi mine (G.R. Allen, Western Australian Museum, personal communication). Another significant recent observation is the proliferation of small mesh (r2v), lightweight Indonesian-made gill nets in the middle and upper Fly. These nets intentionally target small-bodied species, and it is a well-accepted progression that subsistence (and artisanal) fisheries resort to smaller gear 458 A. W. Storey et al. size as a fishery becomes depleted (see discussion in Welcomme, 1985). CPUE may not decline, but the gear targets smaller species, reflecting the loss of slower growing, low-fecundity large-bodied species, and the increase in faster growing, more fecund smaller-bodied species. The appearance of these small mesh nets could be as a result of a cultural transfer of methods from West Papuan refugees coming into the area. However, theory on the adoption of small mesh nets in subsistence/artisanal fisheries is supported by the results of OTML fish catch monitoring data, which show significant declines in fish catch at riverine sites, and now at some oxbow sites. It seems likely that villagers are specifically targeting smaller species out of necessity. Biological monitoring of the Fly River has intentionally targeted fish catch (abundance and biomass) at the request of the State of Papua New Guinea, to ensure a subsistence fishery is maintained for communities downstream of the mine. Analysis of the resultant data (viz. changes in abundance, biomass, species richness, and fish assemblage composition) clearly show the impacts due to the mine. That these impacts are detectable partly reflects the effectiveness of the monitoring approach, but also that the impacts are very significant, and so relatively easy to detect with simple, but robust sampling methods and data analyses. The monitoring program demonstrates the suitability of fish as indicators of ecosystem health, particularly in that they are relatively easy to sample and their mobility and longevity means they can ‘‘integrate’’ temporal and spatial effects at both the catchment scale and at the local scale (Harris, 1995; Harris and Silveira, 1999; Simon, 1999). In recent years monitoring programs for fish (and macroinvertebrates) are moving toward an Index of Biological Integrity (IBI) approach. Originally developed by Karr (1981), it is a multi-metric approach which uses a number of quantifiable fish metrics (parameters) to monitor changes in fish community structure and function. Metrics include species richness, abundance, number of exotic species, fish health, and trophic and habitat guild composition. The information gathered can be broadened by community participation to include data on recreational species (e.g., redfin perch, trout), fish response to short-term fluctuations (spikes) in water quality parameters (e.g., fish kills), and anecdotal information on the historic condition of the rivers and their fish populations. Although detailed knowledge of the life history, habitat requirements, trophic position, and sensitivity to pollutants of all Fly River fish species has not been documented, future adoption of an IBI approach to monitoring fish assemblages in the Fly River may prove productive. In conclusion, over its 25 years of operation, the Ok Tedi Mine has significantly changed the Fly River fishery. Other pressures, such as commercial barramundi fisheries, a growing population, and increased Fish Assemblages in the Fly River System and Effects of the Ok Tedi Copper Mine 459 artisanal fishing with greater access to more efficient nets and boats are also influencing the system. The pressures from exotic species currently in the system, and additional species that may yet be introduced (i.e., carp, tilapia) will further impact the fish fauna. Perhaps the greatest threat in the short term however, is the appearance of ARD along river bank levees, with associated mobilization of metals from mine-derived sediments with a high sulfide content (Bolton et al., 2009; Apte, 2009). ARD is of growing concern for riverine and floodplain fish resources, and will need to be managed carefully from now until after mine closure.

ACKNOWLEDGMENTS

The authors thank all past and present staff of the Environment Department, Ok Tedi Mining Limited, who assisted in the collection of ﬁsh-catch data. Past coordinators of the biological monitoring programs at Ok Tedi are acknowledged for their role in directing the collection of data and inﬂuencing the project: David Balloch (1982–1985), Kent Hortle (1983–1987), Ross Smith (1988–1992), Andrew Storey (1993–1995), Stephen Swales (1996–1998), Charles Tenakanai (1998–2001), and Markson Yarrao (2001–present).

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