Nutrient Cycling- Current Conditions and Impact Assessment

Supplementary Environment Effects Statement, Head Technical Report: Nutrient cycling- current conditions and impact assessment

No. 17 October 2006

Supplementary Environment Effects Statement, Head Technical Report: Nutrient cycling- current conditions and impact assessment

Andrew R. Longmore

October 2006

Department of Primary Industries

Published: Primary Industries Research Victoria, Marine and Freshwater Systems Department of Primary Industries, Queenscliff PO Box 114, Queenscliff, 3225 Victoria

General disclaimer This publication may be of assistance to you but the State of Victoria and its employees do not guarantee that the publication is without flaw of any kind or is wholly appropriate for your particular purposes and therefore disclaims all liability for any error, loss or other consequence which may arise from you relying on any information in this publication. Copyright  The State of Victoria, Department of Primary Industries, 2006.

This publication is copyright. No part may be reproduced by any process except in accordance with the provisions of the Copyright Act 1968. Authorised by the Victorian Government, Bellarine Highway, Queenscliff 3225 Printed by PIRVic Queenscliff, Victoria

Preferred way to cite: Longmore AR (2006) Supplementary Environment Effects Statement, Head Technical Report: Nutrient cycling- current conditions and impact assessment. Marine and Freshwater Systems Report Series No. 17. Primary Industries Research Victoria, Queenscliff.

ISSN 1449-2520 ISBN

SEES Head Technical Report: Nutrient Cycling ii

SEES Head Technical report: Nutrient Cycling iii

Executive Summary

A numerical model developed to describe nutrient cycling processes, the PPBES Integrated Model, indicated that of the 6,000-8,000 tonnes N discharged to the Bay each year, little (~ 700 t) is exported to Bass Strait, or buried in the sediments (~ 1,200 t). There is no long-term build up of nutrients because most of the input is eventually lost from the system as N 2 gas. The process leading to this loss takes place in the sediment, and arises from the coupling of two microbial processes, called nitrification and denitrification. These processes are important because they result in the conversion of nitrogen from forms (ammonium, nitrate) readily available for plant growth, to a form (N 2) that is lost to the atmosphere. While the broad principles of the denitrification process are known, there is still uncertainty about the roles played by various biological communities, including bacteria, infauna and microphytobenthos. Port Phillip Bay appears to be unusually efficient at denitrification, and denitrification has been identified as a key indicator to be maintained. The model predicts that a substantial drop in denitrification efficiency would lead to large increases in plankton production, even at current inputs, and in the extreme cause a shift to a highly enriched state. This is clearly a situation we would wish to avoid. Factors that are thought to influence denitrification include the supply of carbon (including primary production in the water column), the oxygen regime (affected by stratification), type and abundance of infauna and epifauna, presence of microphytobenthos, and the microbial population in the sediment. While denitrification is efficient over much of the Bay floor, there is high spatial and moderate temporal variability. Efficiency is lower in areas close to the major inputs (Hobsons Bay and the Werribee coast) than in the centre of the Bay. There is no evidence that efficiency has changed greatly over the past 10 years. Experiments have demonstrated that a number of species of exotic infauna already present in Port Phillip Bay may affect nutrient cycling and denitrification, but only at densities much higher than currently found in Port Phillip Bay. Microphytobenthos actively take up nutrients from the sediment, and produce oxygen during daylight, which may promote nitrification, but inhibits denitrification. Studies during the Trial Dredge Program did not provide compelling evidence that microphytobenthos plays a significant role in denitrification over most of the Bay floor. Nutrient cycling, and denitrification in particular, may be vulnerable to a number of effects of dredging. Dredging and subsequent dredged material placement has the potential to affect benthic organisms, ambient light, nutrient and oxygen concentrations and phytoplankton growth, all of which may influence nutrient cycling and denitrification. The risks to nutrient cycling presented by dredging must be assessed by the SEES. A formal assessment of the risk of dredging to algal blooms, nutrient cycling and denitrification found that the only bay-wide impact thought possible was that of introducing an exotic species which then caused a bay-wide decline in denitrification. While of major consequence, this chain of events had such a low likelihood that the risk was extremely low. All other impacts applied to one or more of the project areas, rather than the whole of the Bay. The three highest risks to nutrient cycling were all rated at ”low”. The risks included: • the release of labile nutrients from the dredge plume in the Yarra/Hobsons Bay area; • the impact of turbidity on phytoplankton production in the south; • the impact of turbidity on MPB production in the south. All of the impacts are short-term, and restricted to part or all of individual project areas. None are Bay- wide, or likely to last longer than a year after the end of dredging. The mean impact of dredging on nutrient cycling is expected to be the equivalent of a new, one-off input of less than 300 t N. Because none of the risks are medium or higher, no mitigation is proposed for nutrient cycling. Even so, because of the incomplete understanding about specific mechanisms affecting denitrification, and the potentially disastrous impact on the Bay if denitrification fails, monitoring is proposed before, during and after dredging to confirm that high denitrification efficiency is maintained.

SEES Head Technical Report: Nutrient Cycling iv

Table of Contents

Executive Summary ...... iv

1 Introduction...... 1 Study objectives...... 2 Report structure...... 2 Project Description ...... 3

2 Methods ...... 4

3 Policy and Legislation...... 5

4 Factors affecting Nutrient Processes ...... 8 Overview of nutrient cycling and denitrification...... 8 4.01 Primary production and plankton blooms...... 13 4.02 Potential role of fauna in nutrient cycling ...... 14 4.03 Potential role of microphytobenthos in nutrient cycling...... 14 4.04 Climate ...... 15 4.05 Alternative nitrogen recycling processes...... 16 4.06 Measuring nutrient cycling...... 17

5 Nutrient Processes in the Project Areas...... 21 5.1 Yarra River and Hobsons Bay...... 22 Water column processes ...... 23 Benthic processes ...... 28 Summary for the Yarra River and Hobsons Bay ...... 34 5.2 North of the Bay...... 35 Water column processes ...... 36 Benthic processes ...... 40 Summary for north of the Bay...... 45 5.3 South of the Bay...... 46 Water column processes ...... 46 Benthic processes ...... 48 Summary in the south ...... 51 5.4 The Entrance ...... 52

SEES Head Technical report: Nutrient Cycling v

Water column processes...... 53 Benthic processes...... 53 Summary for the Entrance ...... 54 5.5 Summary...... 55 Key nutrient cycling processes...... 55 Key factors affecting processes...... 55 Seasonality...... 56

6 Interactions Arising from the Project...... 57

7 Assessment of Impacts ...... 68 7.0 New evidence of impact...... 68 7.01 ESS studies and conclusions ...... 68 7.02 SEES baseline study ...... 69 7.03 SEES MPB experiments ...... 70 7.04 SEES sediment nutrient bioavailablity experiment ...... 71 7.05 Estimates of nutrient loads from dredging...... 73 7.1 Risk Assessment Method ...... 83 7.2 Yarra River and Hobsons Bay ...... 88 7.7.7.37. 3 North of the Bay ...... 91 7.4 South of the Bay ...... 96 7.5 The Entrance...... 100 7.6 Summary...... 101 Cumulative impacts...... 101 Uncertainty in relation to risk assessment...... 102

8 Mitigation ...... 104

9 Conclusions ...... 105

10 References...... 106

Appendix 1. Channel Deepening Project: Potential effects on Phytoplankton...... 111

SEES Head Technical Report: Nutrient Cycling vi

List of Tables

Table 1 Policy and legislation relevant to nutrient cycling...... 5 Table 2 Estimates of dissolved nutrient load (median ± 95% CI, tonnes ) in Yarra River sediments to be dredged (SKM 2004; Longmore 2006c)...... 29 Table 3 Estimates of bioavailable nutrient load (tonnes, median ± 95% CI) in sediment to be dredged from the Williamstown Channel (SKM 2004; Longmore 2006c)...... 30 Table 4 Total nutrient load (tonnes, mean ± 95%ile) in sediments to be dredged from the Yarra River and Williamstown Channel (Longmore 2006c)...... 31 Table 5 Mean nutrient release over 10 days (tonnes ± 95% CI): change in individual samples averaged and multiplied by dredge volume for each project area and depth interval (Longmore 2006c)...... 31 Table 6 Summary of direct benthic flux measurements ( m mol m -2 d -1) in Hobsons Bay, 1994-2005 (Nicholson et al. 1996; Berelson et al 1998; Longmore 2005)...... 32 Table 7 Estimates of dissolved nutrient load (tonnes, mean ± 95% CI) in sediment to be dredged from the Williamstown Channel (SKM 2004; Longmore 2006c)...... 41 Table 8 Estimates of total nutrient load (tonnes ± 95% CI) to be mobilised during dredging in the Port Melbourne Channel (Longmore 2006c)...... 41 Table 9 Estimates of bioavailable nutrient load released over 10 days (tonnes ± 95% CI) in sediment to be dredged from the Williamstown Channel (Longmore 2004; Longmore 2006c)...... 42 Table 10 Summary of direct benthic flux measurements ( m mol m -2 d -1) in central Port Phillip Bay, 1994- 2005 (Nicholson et al. 1996; Berelson et al . 1998; Longmore 2005)...... 42 Table 11 Estimates of dissolved nutrient load (tonnes, mean ± 95% CI) in sediment to be dredged from the South Channel (SKM 2004; Longmore 2006a; Longmore 2006c)...... 49 Table 12 Total nutrient load to be mobilised by dredging in the South Channel (Longmore 2006c)...... 49 Table 13 Estimates of bioavailable nutrient load (tonnes, mean ± 95% CI).after 10 days in sediment to be dredged from the South Channel (SKM 2004; Longmore 2006c)...... 50 Table 14 Summary of seasonal issues with nutrient cycling...... 56 Table 15 Plausible modes of dredging impact on nutrient cycling...... 57 Table 16 Estimated total nutrient load (mean ± 95% CI) mobilised by dredging in each area (Longmore 2006c) 74 Table 17 Estimates of dissolved nutrient load ( mean ± 95% CI) released to dredge water (Longmore 2006c)...... 74 Table 18 Estimates of bioavailable nutrient load ( mean ± 95% CI) released from particles after settling (Longmore 2006c)...... 75 Table 19 Estimates of bioavailable N load from dredging compared to known existing inputs, fluxes and pools (Harris et al . 1996; *Murray and Parslow 1999; Longmore 2006c)...... 75 Table 20 Estimates of nutrient loading from each dredging activity (Longmore 2006c)...... 81 Table 21 Existing benthic DIN fluxes compared to estimated DIN release from porewater and particles. Dredging loadings are means and 95 th percentile (Longmore 2006c)...... 82 Table 22 Likelihood Guide ...... 83 Table 23 Environmental Consequences Table ...... 84

SEES Head Technical report: Nutrient Cycling vii

Table 24 Indicative dredging schedule for each project area...... 85 Table 25 Summary of bay-wide risks to nutrient cycling...... 87 Table 26 Summary of risks to nutrient cycling in the Yarra River and Hobsons Bay...... 90 Table 27 Summary of risks to nutrient cycling in the north of the Bay...... 94 Table 28 Summary of risks to nutrient cycling in the south of the Bay...... 99 Table 29 Cumulative effects of impacts of dredging on nutrient cycling (Longmore 2004; Longmore 2006c)...... 103

SEES Head Technical Report: Nutrient Cycling viii

List of Figures

Figure 1 Project areas in Port Phillip Bay...... 3 Figure 2 Proportion of annual nitrogen loading to the Bay attributed to known inputs (Harris et al. 1996). 8 Figure 3 Conceptual model of nutrient cycling in Port Phillip Bay (Harris et al . 1996)...... 9 Figure 4 Conceptual model of the links in the N processing chain. Yellow zone is oxic sediment; brown zone is anoxic sediment; black line represents a benthic chamber which traps nutrients leaving the sediment...... 10 Figure 5 Large-scale water circulation formed under northerly (red) and westerly (black) winds. Depth contours are at 5 m intervals (based on Harris et al. 1996)...... 11 Figure 6 Modelled flushing times in Port Phillip Bay (CLT 2006)...... 12 Figure 7 An automated benthic chamber used to directly measure nutrient fluxes between sediment and water column...... 18 Figure 8 Denitrification efficiency measurements (mean and SE) for all studies since 1994. Black – PPBES 1994-96; blue ARC Linkage project 1992-94; green Channel Deepening Project 1994-present; red PPBES and PPB EMP 1994-present ( Berelson et al 1998; Longmore 2005; Ross et al. unpubl)...... 19 Figure 9 Yarra River and Hobsons Bay project area...... 22 Figure 10 Water quality sites sampled for Fisheries Victoria (1990-97, squares) and EPA (1986-present, stars). Note that continuous data is now collected from Central, Hobsons Bay and Long Reef sites for the PPB Environmental Management Plan (Longmore et al . 1996)...... 24 Figure 11 Nitrogen species concentrations in Hobsons Bay, 1990-97 (Longmore et al . 1997)...... 25 Figure 12 Chlorophyll a concentrations from the EPA fixed site in Hobsons Bay, 1989-2005...... 26 Figure 13 Chlorophyll a concentrations at Williamstown, 1990-97 (Longmore et al . 1997)...... 26 Figure 14 Continuous records of dissolved oxygen concentration from near-surface and near-bottom waters in Hobsons Bay, 2002-05 (Longmore 2005)...... 28 Figure 15 Small-scale spatial variability in sediment pore water ammonium concentration: three cores collected 1 m apart in Hobsons Bay (Nicholson et al. 1996)...... 30 Figure 16 Variation in carbon dioxide flux (mean ± 1 SD) in Hobsons Bay, 1994-2006 (Nicholson et al. 1996; Berelson et al 1998; Longmore 2005)...... 33 Figure 17 Variation in denitrification efficiency (mean ± 1 SD) in Hobsons Bay, 1994-2006 (Nicholson et al. 1996; Berelson et al 1998; Longmore 2005)...... 33 Figure 18 Northern project area...... 35 Figure 19 Surface water nutrient concentrations, interpolated from 12 monthly underway surveys in 1993 (derived from Shao and Fox 1996)...... 36 Figure 20 Nitrogen species concentrations in northern Port Phillip Bay, 1990-97 (Longmore et al . 1997). 37 Figure 21 Annual N load to the Bay from the Western Treatment Plant (Trevor Gulovsen, Melbourne Water, and Melbourne Water Annual Social and Environmental Reports)...... 38 Figure 22 Chlorophyll a concentrations from the EPA Central site, 1989-2005...... 39 Figure 23 Chlorophyll measurements at the Central site, early 2005: Yarra flood response (Longmore 2005). 40

SEES Head Technical report: Nutrient Cycling ix

Figure 24 Variation in carbon dioxide flux (mean ± 1 SD) in central Port Phillip Bay, 1994-2006 (Nicholson et al. 1996; Berelson et al . 1998; Longmore 2005)...... 43 Figure 25 Variation in denitrification efficiency (mean ± 1 SD) in central Port Phillip Bay, 1994-2006 (Nicholson et al. 1996; Berelson et al . 1998; Longmore 2005)...... 43 Figure 26 Southern project area...... 46 Figure 27 Nitrogen species concentration in southern Port Phillip Bay, 1990-97 (Longmore et al . 1997). 47 Figure 28 Chlorophyll a concentrations from Blairgowrie, 1990-97 (Longmore et al . 1997)...... 48 Figure 29 The Entrance project area...... 52 Figure 30 Nitrogen species concentrations at the Entrance, 1994-95 (Longmore et al . 1996)...... 53 Figure 31 Removal of seabed by dredge prevents denitrification. Symbols courtesy of the Integration and Application Network ( ian.umces.edu/symbols/ ), University of Maryland Center for Environmental Science. 58 Figure 32 Nutrients from the dredge plume increase phytoplankton production...... 59 Figure 33 Nutrients from the dredge plume cause toxic algal blooms...... 60 Figure 34 Turbidity from the dredge plume reduces phytoplankton production...... 61 Figure 35 Turbidity from the dredge plume reduces microphytobenthos production...... 62 Figure 36 Sediment from the dredge plume buries microphytobenthos...... 63 Figure 37 Sediment from the dredge plume buries infauna...... 64 Figure 38 Toxicants from the dredge plume reduce zooplankton grazing on phytoplankton...... 65 Figure 39 Toxicants from the dredge plume affect infauna...... 66 Figure 40 Invasion by exotic species alter infauna...... 67 Figure 41 Ammonium concentrations in elutriates from each area (Longmore 2006c)...... 72 Figure 42 Release of ammonium from sediment particles over 10 days, after suspension in seawater (Longmore 2006c)...... 73 Figure 43 Contours for the fraction of time suspended solids exceed 2 mg L -1 in the dredge plume during dredging of contaminated (top LHS) and uncontaminated (top RHS) sediment in the Yarra River and Williamstown Channel (bottom RHS); CLT 2006...... 77 Figure 44 Contours for the fraction of time suspended solids exceed 2 mg L -1 in the dredge plume during dredging in the Port Melbourne Channel (CLT 2006)...... 78 Figure 45 Contours for the fraction of time suspended solids exceed 2 mg L -1 in the dredge plume during dredging in the South Channel east of Hovell Pile by the Cornelis Zanen (top LHS) and the Queen (top RHS), west of Hovell Pile (bottom LHS) and in South Channel West (bottom RHS); CLT 2006. 79

SEES Head Technical Report: Nutrient Cycling x

1 Introduction

The Port of Melbourne Corporation (PoMC) is responsible for establishing, dredging and maintaining shipping channels in port waters in Melbourne and Geelong. In order to improve commercial shipping access to the Port of Melbourne, the Port of Melbourne Corporation (PoMC) proposes to deepen the shipping channels in Port Phillip Bay (the Bay) to the Port of Melbourne (PoM) to a depth sufficient to accommodate ships of 14 metre draught, known as the Channel Deepening Project. Because of the potential environmental impacts that may arise from dredging, the Government directed PoMC to carry out an Environmental Effects Statement (EES) as required under the Environmental Effects Act 1978. An EES for the Channel Deepening Project was released for public comment on 5 July 2004 and an Independent Panel Hearing was conducted from 21 September 2004 through to 17 December 2004. A statement from the Minister for Planning was released publicly on 31 March 2005, together with the Independent Panel’s report. The Minister for Planning’s statement included a directive that a Supplementary Environment Effects Statement (SEES) be prepared by the PoMC. This included a re- assessment of the potential risks to nutrient cycling posed by dredging. The SEES Assessment Guidelines (DSE 2005, Section 6.3) stated that a nutrient cycling study was required, to carry out “Further assessment of key environmental effects and risks to demonstrate the ability of the proposed dredging campaign to avoid or effectively minimise adverse effects in order to comply with policy requirements: ……………. vi) Significant risk to human heath or ecological health from blooms of toxic algae species or resulting from short-term nutrient release; vii) Significant risk to ecological health from altered ecosystem nutrient budgets or a protracted reduction in the denitrification capacity of Bay sediments. Potentially significant risks need to be addressed in the SEES, by demonstrating that the risk would be acceptably low either without special measures being implemented (i.e. the project is unlikely to have a significant effect) or with special measures being implemented (i.e. through proposed risk avoidance or mitigation or other management measures).” The SEES is further required to: • Use refined turbidity modelling of the dredging campaign to assess the effects of turbidity plumes and sediment smothering on primary production of key marine plants and communities (including seagrasses, kelp and microphytobenthos), denitrification and other relevant ecological processes; • Characterise risks to the denitrification process arising from the proposed dredging campaign due to both smothering of sediments and release of pore water. Consideration should be given to the cumulative effect of other spatial and temporal influences on nitrogen cycling in the Bay, including storm flows from the Bay’s catchment and the effects of invasive introduced marine species; • Characterise both the risk of algal blooms occurring as a result of dredging, the driving factors and the consequent risks to ecological processes and particular communities or species in the Bay. Primary reliance may be placed on synthesis of currently available knowledge. The development of a strategy to manage the risk of algal blooms should also be based on this synthesis; • Identify the means by which the release of nutrients from sediments will be minimised through the design and management of the dredging program; • Identify and fully justify relevant threshold limits and performance criteria (for different areas) for minimising the effects of dredging on both primary production and denitrification, both in terms of their scientific basis and their achievability using best practice approaches; and • Provide reliable estimates of nitrogen inputs from both the proposed capital dredging, and outline any proposed arrangements to offset these nitrogen inputs.

SEES Head Technical report: Nutrient Cycling 1

The report was also required to have regard for the general guidance outlined in the EES Guidelines for ecological processes and nutrient cycling. The EES Guidelines required an assessment of the effects of the proposed dredging and management of dredged material on: • Potential changes in the biogeochemical activity of seafloor and estuarine sediments; • Potential impacts on oxygen irrigation of sediments and nutrient cycling (especially of nitrogen), including in intertidal environments; and • Potential resuspension of algal cysts and stimulation of algal growth, including toxic phytoplankton, and their potential impacts. All of these issues are addressed in this report, except discussion of marine plants other than phytoplankton and microphytobenthos, which may be found in the Marine Ecology Head Report.. Study objectives

The SEES Assessment Guidelines provided objectives for the SEES, including: ………………… 3). To avoid unacceptable adverse long-term effects, and to minimise any short-term adverse effects, on the Bay’s ecological processes and species or areas of conservation significance, as well as on the Bay’s fringes, the area immediately outside Port Phillip Heads and the Yarra estuary, taking account of other long-term environmental influences; 4) To avoid adverse effects on human health and minimise any short-term adverse effects on public amenity, as a result of diminished water quality (e.g. turbidity, algae or toxicants) and dredging activities, in particular around sections of the foreshore and waters regularly used for recreation, tourism and other purposes.

The objectives of this report are to address the requirements of the EES and SEES Guidelines, by: • describing the current condition of nutrient cycling in Port Phillip Bay; • describing credible pathways by which dredging may affect nutrient cycling; • carrying out a formal assessment of the risks for each pathway and area of the Bay, distinguishing between long-term, irreversible changes and small-scale transient changes; and • recommending mitigation where necessary.

Report structure

This report comprises two main parts: Chapters 1-5, which provide an assessment of current conditions in Port Phillip Bay, and Chapters 6-9, which provide an assessment of the potential impacts of dredging on nutrient cycling and denitrification. Current conditions and risk assessments are provided for four main Project Areas (see below). Nutrient cycling is discussed in relation to measurements in the water column, in the sediments, and of the transfer between sediment and water column (called “benthic fluxes”) in each Project Area. The report includes extensive discussion of phytoplankton, because of the key role it plays in nutrient cycling.

SEES Head Technical Report: Nutrient Cycling 2

Project Description

To accommodate shipping with increased draught, the Channel Deepening Project proposes dredging in the Entrance, parts of the South Channel, the Port Melbourne Channel, Williamstown Channel, the Yarra River, and in various berths. Dredged material will be placed in an existing Dredge Material Ground (the Port of Melbourne DMG) and a new ground in the south-east (the SE DMG). The Bay has been divided into four Project Areas: the Yarra River and Hobsons Bay; the northern half of Port Phillip Bay; the southern half of Port Phillip Bay; and the Entrance (Fig 1).

Figure 1 Project areas in Port Phillip Bay.

SEES Head Technical report: Nutrient Cycling 3

2 Methods

Specific steps taken in preparing the first part of this report include: • Review of Channel Deepening Project EES documentation, particularly the Water Quality and Sediment Chemistry reports. • Review of information from the Port Phillip Bay Environmental Study • Review and incorporation of relevant recent international literature, including that on alternative N cycling pathways to denitrification. • Review and incorporation of unpublished research, including that carried out on the impact of marine pests on nutrient cycling by Melbourne University, the Department of Sustainability and Environment and the Department of Primary Industries. • Review and incorporation of all relevant Trial Dredge Program reports. • Incorporation of baseline denitrification monitoring carried out since the EES. • Review and incorporation of paper on risk and management of algal blooms. • Production of a revised nutrient cycling “current conditions” report.

Steps taken in preparing the impact assessment include:

• Incorporation of findings from post-TDP studies on sediment nutrient content and bioavailability, and recalculation of nutrient loads to the Bay from dredging. • Carrying out of assessment of impacts, according to the revised risk management framework and SEES guidelines, Bay-wide and in four smaller areas of the Bay. Such impacts are assessed against (i) “natural” variability, including denitrification as a key process vulnerable to impact; (ii) project evaluation objectives listed in the assessment guidelines, and (iii) applicable policy and legislation. • Recommendation of mitigation options (including suitable indicators, trigger points and thresholds, if necessary). • Production of a “current conditions and impact assessment” report (this report).

SEES Head Technical Report: Nutrient Cycling 4

3 Policy and Legislation

A number of state and federal government policies and management plans have a bearing on nutrient cycling and dredging in Port Phillip Bay, either directly or incidentally (Table 1).

Table 1 Policy and legislation relevant to nutrient cycling.

Legislature Name of policy/act/guideline

State (EPA) State Environment Protection Policy State (DSE) Port Phillip Bay Environmental Management Plan State (DPI) Victorian Shellfish Quality Assurance Plan Federal ANZECC Water Quality Guidelines State (EPA) Best Practice Environmental Management Guidelines for Dredging Federal Commonwealth Quarantine Act

3.1 State Environment Protection Policy (SEPP)

The Victorian Government’s aspirations to protect the long-term health of the state’s waters is expressed in the State Environment Protection Policy (SEPP), which provides a policy framework for the protection of water quality, and is proclaimed under Section 16(2) of the Environment Protection Act 1970. Protection of waters in Port Phillip Bay is specified in SEPP (Waters of Victoria) Schedule F6 Waters of Port Phillip Bay (1997). Requirements regarding dredging and disposal of dredged material in Port Phillip Bay are outlined in Clause 13, which states that: ‘Protection agencies or bodies undertaking dredging or spoil disposal must ensure that: • these activities are conducted in accordance with the current best practice or any code of best practice approved by the Authority; • these activities are conducted and managed to ensure local exceedances of the prescribed environmental objectives are confined to the smallest practicable area and over the shortest practicable time in the vicinity of the dredging and disposal operation; • these activities do not re-suspend and/or disperse sediments or accumulated contaminants that will be detrimental to the long term protection of beneficial uses; and • dredge spoil is disposed to land in preference to water wherever practicable and environmentally beneficial as determined by the Authority.’

SEPP Schedule F6 Waters of Port Phillip Bay (1997) divides Port Phillip Bay into six segments on the basis of differing beneficial uses and/or location, each with defined environmental quality objectives. The Hobsons segment includes the Port areas, the Yarra mouth and Hobsons Bay. The Werribee segment surrounds the Western Treatment Plant outfalls. The Corio segment includes Corio Bay; the Inshore segment covers waters within 600 m of the shore; Aquatic Reserves cover areas with statutory protection; and the General Segment covers all other areas (the majority of the Bay). Protection of waters of the Yarra Catchment is specified in SEPP (Waters of Victoria) Schedule F7 Waters of the Yarra Catchment (1999). SEPP Schedule F7 divides the Yarra Catchment into segments, each with

SEES Head Technical report: Nutrient Cycling 5

defined environmental quality objectives. Of particular relevance to the Port of Melbourne is the Yarra Port segment (SEPP 1999). Dredging is planned in the Yarra Port section (covered by schedule F7) and also in Hobson and General segments (covered by Schedule F6), but dredging plumes could potentially also affect Inshore and Aquatic Reserve segments. The SEPPs define beneficial uses, and provide targets for specific indicators. These do not include nutrient concentrations or any measure of nutrient cycling, but rely on chlorophyll a, an indicator of the amount of microscopic plant abundance, which is a likely indicator of excessive nutrient supply. The Policy states that inputs of nutrients must be below levels defined by the Authority as posing an unacceptable risk to the chlorophyll objectives. Progress toward the achievement of SEPP objectives is measured in terms of water quality and nutrient status (EPA 2002), but not specifically in terms of nutrient cycling. Even so, the decision to implement a 1,000 tonne reduction in annual nitrogen load to the Bay (EPA 1997) was made as a precautionary measure to protect nitrogen cycling.

3.2 Port Phillip Bay Environmental Management Plan (PPBEMP)

The SEPP revision in 1997 foreshadowed the development of an Environmental Management Plan (EMP). The EMP was the State Government’s response to the recommendations arising from the CSIRO Port Phillip Bay Environmental Study. Its purpose was to provide the framework within which the various authorities involved in managing Port Phillip Bay could work together to protect the Bay, by developing a shared understanding of environmental objectives for the Bay, and key risks to achieving the objectives. The EMP included the key environmental objective “to conserve biodiversity”, and recognised as key risks (DNRE 2002): • deterioration of water quality, • increased nutrient loading and detrimental changes to nutrient cycling, • increased suspended solids levels, • exotic marine pests and • physical disturbance of habitats. To achieve its objectives, the EMP called for a reduction in annual nitrogen inputs of 1,000 tonnes by 2006, and set a sub-objective to ensure there are no net additions of nitrogen to the Bay. The reduction is to be achieved from the Western Treatment Plant (500 t) and rivers and streams (500 t). The EMP also established a Bay nitrogen cycling monitoring program, with the objective: “To detect, as early as possible with current scientific understanding, detrimental changes to critical elements of Bay nitrogen cycling processes that indicate an increased risk of eutrophication at Bay-wide and regional scales” . The monitoring program was developed through a series of expert workshops and commissioned reports (DNRE 2002). The current report draws heavily on this monitoring program.

3.3 Victorian Shellfish Quality Assurance Plan (VSQAP)

This plan outlines the monitoring needed to enable shellfish producers to gain export accreditation. Phytoplankton and bacteria are monitored every fortnight at three growing zones in Port Phillip Bay (Clifton Springs, Grassy Point and Dromana). Toxicity tests are carried out if nuisance phytoplankton reach prescribed levels, and harvesting is halted if phytoplankton are toxic, or if bacteria reach prescribed levels. Bacteria are generally transported to shellfish growing areas by storm waters, and predictive models have been developed to allow rainfall to be used as a trigger to close growing areas. VSQAP is relevant to this report because algal blooms reflect a change in primary production, which influences nutrient cycling (see below).

SEES Head Technical Report: Nutrient Cycling 6

3.4 ANZECC Water Quality Guidelines

ANZECC recently revised its guidelines for fresh and marine water quality (ANZECC/ARMCANZ 2000). Previous water quality guidelines dealt with marine and freshwater systems by applying single values to indicators without accounting for other factors that may reduce or exacerbate their effects. With increased understanding of the complexity of ecosystems, imposing uniform goals across different locations is now considered unrealistic. The revised guidelines reflect a new approach, which is broader and more holistic. Regional nutrient trigger values are derived from reference sites, and then applied to other sites that may be affected by human activity (ANZECC/ARMCANZ 2000). They are intended to trigger further investigation if they are exceeded. They may also be compared to national default trigger values, but there is abundant evidence that these are of limited use in Port Phillip Bay (too low for phosphorus forms, and too high for nitrogen forms; EPA 2002). No triggers or other advice concerning nutrient cycling is provided in the guidelines, in part because there have been relatively few studies on nutrient cycling in Australian marine and estuarine waters.

3.5 Best Practice Environmental Management Guidelines for Dredging

Guidelines to advise on environmental requirements for dredging and disposal of sediments in Victoria were developed firstly as the Trial Dredge Protocol (EPA 1992) and then as the Best Practice Environmental Management Guidelines for Dredging (BPEMGD) (EPA 2001). These guidelines apply to both dredging and dredged material disposal within Victoria. Disposal of dredged material on the open coast, other than for beach renourishment, must additionally comply with the requirements of the Commonwealth Environment Protection (Sea Dumping) Act. The guidelines focus principally on preventing toxicity, and minimising impacts on water quality, but also flag the potential of dredging to impact denitrification. The guidelines recommend monitoring of nutrient levels during dredging. Elutriate tests are required when there is insufficient information to demonstrate that relevant water quality criteria (ANZECC/ARMCANZ 2000) will not be exceeded after allowing for mixing that occurs within 4 hours of dumping (EPA 2001). Such tests are used in this report to assess the initial release of nutrients from dredged material.

3.6 Commonwealth Quarantine Act

The plants and animals living in the Bay have a role in nutrient cycling. To that extent there is a need to protect the biota from exotic species that may impact and disrupt the nutrient cycling. The Quarantine Act is important for the controls it places on the management of ballast water, which is the major vector for the introduction of exotic species to the Bay. The Act operates by applying a risk assessment to all commercial shipping entering Australian waters, and providing for generic treatment procedures for all ballast waters classed as high risk, that are intended for discharge in Australian waters.

SEES Head Technical report: Nutrient Cycling 7

4 Factors affecting Nutrient Processes

The purpose of this chapter is to provide a summary of nutrient cycling processes as they apply to Port Phillip Bay, and those physical and biological factors that may affect nutrient cycling. Overview of nutrient cycling and denitrification The food supply of all animals in the Bay (fauna on and in the sediment, fish, shellfish, penguins, seabirds and dolphins) ultimately depends on the production of plants, and plant growth is affected by nutrient supply. Too little nutrient may lead to restricted growth. Too much nutrient may lead to the explosive growth of a family or species of plant, and a range of undesirable impacts that may follow, including aesthetic, ecosystem and human health impacts. How nutrients are transferred through a system is referred to as nutrient cycling. Nitrogen has been identified as a key nutrient in PPB, with a major influence on plant growth. Each year, 6,000-9,500 tonnes of nitrogen (N) enters Port Phillip Bay (Parslow et al . 1999). Small amounts of N come from the atmosphere (13%; Fig 2), large N loads come from the Western Treatment Plant (46%) and large, sporadic N loads come from the Yarra (24%) and the Werribee and Patterson Rivers, Mordialloc Creek and smaller creeks and drains (17%) (Harris et al . 1996). It is the sporadic N loads from rivers, and to a lesser extent from the Western Treatment Plant, that creates great variability in annual N loads. In wet years, the annual N loads are higher than dry years, when there are lower loads from the rivers, creeks and drains. Nevertheless, despite the wide natural variations, Port Phillip Bay remains resilient in years of low and high N loads. There is no evidence of a significant change in water quality in the Bay over more than 35 years (Longmore 1992; Harris et al. 1996).

24% Yarra/Maribyrnong R.

46% Hobsons Bay

Western Treatment Plant 17% 15 m 20 m

Other rivers and streams 13% Atmosphere

20 m Great Sands 0 15 km Bass Strait

Figure 2 Proportion of annual nitrogen loading to the Bay attributed to known inputs (Harris et al. 1996).

SEES Head Technical Report: Nutrient Cycling 8

Nitrogen is important for plant growth, including phytoplankton, which is the basis of the food chain in Port Phillip Bay (Harris et al . 1996). The receiving environments of the two largest sources of N loads – off Werribee and in Hobsons Bay- are, not surprisingly, two of the most highly productive areas of Port Phillip Bay, as indicated by phytoplankton biomass. The Port Phillip Bay Environmental Study (PPBES) was conducted for Melbourne Water by CSIRO in 1992-1996 in order to assess the impact of the Western Treatment Plant at Werribee on Port Phillip Bay. The PPBES Integrated Model predicts that a substantial drop in denitrification efficiency would lead to large increases in plankton production, and in the extreme cause a shift to a highly enriched state potentially leading to an increased frequency of algal blooms (Murray and Parslow 1997). It concluded that Port Phillip Bay could process 11,000 to 17,000 tonnes N a year, approximately double the current annual N load from the catchment, without irreversible damage. To implement the Port Phillip Bay Environmental Management Plan, Melbourne Water reduced the annual N load from the Western Treatment Plant by 500 tonnes N from 2005, and a similar reduction is expected from the rivers (Yarra-Maribyrnong 350t, Patterson R. 150t; Fletcher and Deletic 2006) over a longer period. Though plankton biomass and ambient nutrient concentrations in Port Phillip Bay are low, nutrient demand by the phytoplankton and microphytobenthos is high (about 6 times the external input). This implies intense recycling of nutrients (Fig 3). The PPBES Integrated Model assigned very little of the 6,000-9,500 tonnes N discharged to the Bay each year to losses to Bass Strait (~ 700 t). Similarly, the model assumed that there was only a small long-term build up of nutrients in the sediments (~1,200 t) and that virtually all of rest of the input (around 5,000 tonnes N) is eventually lost from the system as N 2 gas. The process leading to this loss takes place in the sediment, and arises from the coupling of two microbial processes, called nitrification and denitrification .

WATER Zooplankton Phytoplankton Nutrients

Diatoms

Issues: N2 MPB

C106 N16 P 1Si 17 O2 CO 2 + NH 3/NO 3 + PO 4 + SiO 4 Bacteria & Deposit Feeders ?

Burial Bioturbators & Bioirrigators SEDIMENT

Figure 3 Conceptual model of nutrient cycling in Port Phillip Bay (Harris et al . 1996).

Organic nitrogen in the sediment, derived mainly from decaying phytoplankton, is first converted to ammonium (Fig 4). Nitrification involves the microbial conversion of ammonium to nitrate in the presence of oxygen. Denitrification involves the conversion, by a different suite of microbes, of nitrate to N2 gas. These processes are important because they result in the conversion of nitrogen from forms

SEES Head Technical report: Nutrient Cycling 9

readily available for plant growth (ammonium, nitrate), to a form that is lost to the atmosphere (N 2). A major outcome of the PPBES was the realisation that the resilience of Bay ecosystems to nutrient inputs depended strongly on an efficient coupling of nitrification and denitrification processes, which effectively removes nitrogen from the Bay to the atmosphere (Harris et al . 1996). Efficient coupling depends on the existence of oxic and anoxic zones in the sediment, and effective means of transporting nutrients between the zones (Fig 4). Sediment that is uniformly anoxic cannot nitrify, while sediment that is uniformly oxic cannot denitrify. The oxygen regime in the sediment is therefore a strong determinant of denitrification, as are the rates of supply of oxygen from the atmosphere and organic matter from primary producers (Murray and Parslow 1997). Processes that enhance advection in the sediment (bioturbation and bio- irrigation by infauna) are also important. Fauna, and possibly microphytobenthos, may play a role in the distribution of oxic and anoxic zones and in transporting nutrients between the zones.

Organic NH NO N matter 4 Oxygen 3 2

Oxic

Ammonium Nitrate N2 Nitrification

Deamination Denitrification

Figure 4 Conceptual model of the links in the N processing chain. Yellow zone is oxic sediment; brown zone is anoxic sediment; black line represents a benthic chamber which traps nutrients leaving the sediment. Limited experimental evidence available for the Bay indicates that increasing nutrient loads should lead to a decline in denitrification efficiency (the relative proportion of N lost as N 2, compared to ammonium and nitrate; Murray and Parslow 1997). Efficiency is already lower in areas close to the major inputs (Hobsons Bay and the Werribee coast) than in the centre of the Bay (Berelson et al . 1998). By contrast, decreasing N loads, as has occurred with the load reduction from the Western Treatment Plant, should lead to an increase in denitrification efficiency. The PPBES model indicated that even at current N loads, a complete loss of denitrification capacity would lead to eutrophic conditions (Parslow and Murray 1999). Protection of dentrification capacity is therefore a key need if we wish to maintain water quality in the Bay.

Nutrient cycling, and denitrification in particular, may be vulnerable to a number of effects of dredging. Dredging and subsequent dredged material placement has the potential to affect benthic organisms,

SEES Head Technical Report: Nutrient Cycling 10

ambient light, nutrient and oxygen concentrations and phytoplankton growth, all of which may influence nutrient cycling and denitrification. The SEES must assess the risks to nutrient cycling presented by dredging. This report provides the background against which such an assessment may be made, and the results of the assessment.

The way in which nutrients are processed depends on a number of physical, chemical and biological factors. Physical factors include: • water circulation (tidal or wind-driven: Fig 5): tidal currents dominate on and south of the Sands, but the net movement of water around the rest of the Bay is wind-driven , • evaporation and stream flow, which nearly balance, so that except near discharges, salinity is close to oceanic; • residence times: the theoretical residence time for water in the centre of the Bay is about one year (Walker 1999), which is long enough for nutrients entering the Bay to be taken up and recycled through the plankton many times before they could be flushed to Bass Strait. Flushing times are much shorter on and south of the Sands (Fig 6).

Figure 5 Large-scale water circulation formed under northerly (red) and westerly (black) winds. Depth contours are at 5 m intervals (based on Harris et al. 1996).

SEES Head Technical report: Nutrient Cycling 11

The Yarra/Maribyrnong discharge typically flows down the eastern coast. We could therefore expect that the increased plankton growth, and ultimately increased cycling of nutrients from settled planktonic matter, that will arise from nutrient inputs, might reflect these spatial patterns. Depending on the strength of the circulation, the impacts of nutrient inputs may occur some distance from the location of the input. In addition to the physical setting, climatic factors may also play a role in nutrient cycling. The potential impacts of dredging on nutrient cycling may only be understood if the “natural” seasonal and spatial variations listed above are also understood. Chemical factors include the types, concentrations and rates of supply of various nutrients, particularly those forms available for plant growth. Biological factors include the types of plants and animals involved in either the production or consumption of plant matter. One of the key processes for determining nutrient cycling is the transformation of nutrients through the plankton, and supply of organic matter to the sediment. Other processes that may be important include the role of fauna in the sediment, and the role of microphytobenthos on the sediment surface. These are all discussed below.

Figure 6 Modelled flushing times in Port Phillip Bay (CLT 2006).

SEES Head Technical Report: Nutrient Cycling 12

4.01 Primary production and plankton blooms

Features within the water column important to nutrient cycling include measures of, and interactions between, nutrients, light, temperature and biota (animals and plants). One of the most important components for measurement is the amount of free-floating algae (phytoplankton), as excessive amounts of phytoplankton (algal blooms) can lead to serious ecological and aesthetic issues. There are two measures of algal growth that need to be considered. One is a measure of production/productivity and provides information on how quickly the phytoplankton is growing. The other measure is biomass, which provides information on how much of the phytoplankton is present. Biomass will be controlled by productivity, as well as by the rate at which the phytoplankton is removed (eg through grazing or death). Phytoplankton can be broadly placed into two groups: those that form part of the normal food web, and those (often non-native) that may be toxic to one or more consumers. Non-toxic algal blooms

A review of phytoplankton ecology appears in Appendix 1. Phytoplankton growth is generally controlled by physico-chemical factors, including light, nutrient availability and temperature (Wood and Beardall 1992). A bloom may be defined as a rapid increase in the number of phytoplankton cells. Blooms occur naturally, even in pristine systems, but their size, intensity, frequency and duration may all be affected by elevated nutrient inputs. Non-toxic blooms may impact on the health of the Bay in a number of ways, including fish kills as a result of oxygen depletion, and as a source of organic carbon to alter nutrient cycling. The former has never been recorded in Port Phillip Bay. The latter is the basis of our understanding of the differences in nutrient cycling between Hobsons Bay and the rest of Port Phillip Bay. Hobsons Bay is the site of most of the blooms observed in Port Phillip Bay (Arnott et al. 1997). Murray and Parslow (1997) concluded that phytoplankton in Port Phillip Bay are more likely to be nutrient-limited than light limited in shallow waters (< 10 m), but that light limitation was possible in the deeper central waters. Beardall et al . (1997) considered that the smaller cells (picoplankton, < 2 µm diameter) might be more light-limited than the larger classes. They found that throughout the entire water column during daylight, there was insufficient light 80% of the time to allow maximum phytoplankton growth. Conversely, light above the saturation level (~ 100-200 µEi m -2 s -1) occurs in the top 2-3 m of the water column for much of the day during summer, and at the surface on most days throughout the year. Thus, it is not correct to conclude that phytoplankton growth everywhere is light- limited 80% of the time. For example, during a typical midday in summer, saturating light reaches a depth of 13 m south of the Sands, and a depth of 3-4 m in Hobsons Bay. Further, in the well-mixed waters of most of Port Phillip Bay, plankton may well be transported from depth into the light-saturated area several times per day. Growth is possible below the light saturation level, but it will not be at the maximum rate. The work of Beardall et al . (1997) serves to remind us that we cannot assume nutrients are the only things that may limit phytoplankton growth in the Bay. Beattie et al . (1997) found increased plankton growth in most months with added ammonium, when cells were incubated at or near the saturating light level. Added nitrogen did not enhance growth at temperatures < 12 oC, indicating that light-saturated growth is a function of both temperature and limiting nutrient. At temperatures above 14 oC, there was a linear relationship between increased biomass and ammonium consumed, with the biomass increase averaging 1-2.5 µg chlorophyll a per µM ammonium consumed. The experiments of Beattie et al. (1997) indicated that a rapid increase in growth (particularly of the picoplankton) of 2-3 doublings per day in chlorophyll could be expected if nitrogen was added in summer, whereas the response could be much slower (< 0.5 doublings per day) in winter. If the new cells are small, they may well be grazed by microzooplankton, but if they are large (> 20 µm), they are much more likely to sediment than to be grazed. The usual spring-summer increase in chlorophyll concentration is at the optimal time for secondary producers (eg. mussels, scallops), when their metabolic requirements are highest due to high temperature and spawning. A comparison of plankton biomass to production indicates that grazing is intense in spring, summer and autumn, but much less in winter (Beardall et al. 1997).

SEES Head Technical report: Nutrient Cycling 13

Toxic algal blooms

Some algae produce compounds that may be toxic to fauna and humans (see Appendix 1). Symptoms caused by different classes of algae include paralysis, amnesia, diarrhoea, neurologic effects and bitter taste. Toxic algal blooms have been a problem both in Hobsons Bay (with dinoflagellates) and over the Bay as a whole (with the diatom Rhizosolenia cf chunii ). Dinoflagellates are responsible for Paralytic Shellfish Poisoning (PSP). The toxin accumulates in shellfish, and is thereby transferred to human consumers of the shellfish. The cysts of PSP-forming dinoflagellates may also be toxic. Dinoflagellates are also responsible for Diarrhetic Shellfish Poisoning (DSP), which is also transferred to humans via shellfish. Amnesic Shellfish Poisoning (ASP), caused by yet another group of dinoflagellates, also accumulates in shellfish, as does Neurotoxic Shellfish Poisoning (NSP). The NSP toxin is also toxic to fish. While very broad conditions leading to toxic algal blooms are known (calm, sunny weather after high flows in the Yarra in summer), there have been no studies which provide more specific trigger conditions in the Bay. Many diatom species also form cysts that are an important part of their life cycle, acting as seed reservoirs to maintain the population. Blooms of the diatom Rhizosolenia cf. chunii have occurred intermittently throughout much of Port Phillip Bay in recent years. When this diatom is ingested in large quantities by filter feeding bivalve shellfish such as mussels and oysters, it imparts an intensely bitter and persistent taste to the shellfish rendering them unsuitable for human consumption. High shellfish mortality may also occur (Parry et al. 1989; Nicholson et al. 1998).

4.02 Potential role of fauna in nutrient cycling

Two issues are relevant to nutrient cycling. The first is the role of native fauna, and the second is the potential impact of exotic fauna. Existing data on the role of fauna in denitrification are inconclusive. International studies have shown that bio-irrigation may lead to increases or decreases in denitrification efficiency, depending on size, number and spacing of burrows, oxygen content of water and rate of irrigation (Aller and Aller 1992; Maye r et al. 1995). Local studies have indicated that bioirrigation by fauna enhances nutrient fluxes from the sediment (Nicholson et al. 1996; Bird et al . 1999), but there is only a weak association of denitrification with bioirrigation in Port Phillip Bay (Berelson et al. 1998, 1999). Similarly, the evidence for dramatic impact of exotic fauna (the Mediterranean fanworm Sabella spallanzanii , and Northern Pacific Seastar, Asterias amurensis ) on nutrient cycling at current densities is lacking (Longmore et al . 1996a; Ross et al . unpublished). Conversely, Macreadie (2004) used laboratory incubations of sediment cores to examine the impact of an introduced bivalve, Raeta pulchella , on denitrification efficiency. Macreadie found that denitrification efficiency declined dramatically with increasing bivalve density over the range of densities found in northern Port Phillip Bay, from about 70% to near zero, but such an impact has not yet been demonstrated in the field. The potential role of dredging impact on nutrient cycling via effects on infauna may therefore range from minimal to highly significant.

4.03 Potential role of microphytobenthos in nutrient cycling

Microphytobenthos (MPB) are single-celled microscopic plants living at the sediment surface. MPB have the potential to impact on denitrification rates through two processes: i) oxygen production and consumption, and ii) nitrogen recycling. As oxygen-producing plants, MPB may alter the oxygen regime in and above the sediment surface on a daily basis, which may affect the capacity of the sediment to nitrify and /or denitrify. Modelling (Harris et al. 1996) indicated the key role MPB may play in the Bay nitrogen cycle. The MPB has a low biomass (nitrogen pool of about 400 tonnes in Port Phillip Bay), but turns over about 50 times per year. Living on

SEES Head Technical Report: Nutrient Cycling 14

and near the sediment surface, the MPB intercepts nutrients remineralised within the sediments that would otherwise diffuse into the water column, or to sites for nitrification/denitrification. A comparison of benthic fluxes to external nitrogen inputs to the Bay suggests that nitrogen is recycled several times through the sediment surface before being lost from the system as N 2 gas. On each pass through the remineralisation loop, about 35% is denitrified, 20% escapes to the water column, and 45% is taken up by MPB. Loss of MPB, or a decline in photosynthesis, may therefore have a significant impact on denitrification efficiency. If the loss leads to increased nutrient flux to the water column, denitrification efficiency will be reduced. Conversely, if the loss leads to increased nutrient flux to the nitrification/denitrification sites in the sediment, denitrification efficiency will be increased. Infauna may increase the flux of nutrients from the sediment into the overlying water, and increase the productivity of MPB (Swanberg 1991). Alternatively, flushing nutrients from the sediment may reduce the access of MPB to nutrients in the sediment. Oxygen injected deep in the sediment by bio-irrigation may enhance nitrification, leading to the possibility of enhanced denitrification. Bioturbation by infauna may reduce MPB productivity by mixing MPB into the sediment, reducing exposure to light (Sundback et al . 2003).

4.04 Climate

Many biological processes double in rate for an increase of about 10 oC in temperature. Since the annual temperature cycle in Port Phillip Bay is about 10 oC, we could expect seasonal differences in biological processes such as plankton growth, and bacterial processes critical to nutrient cycling. The annual light cycle also impacts on plant growth, as noted above. In Port Philip Bay, winter production of phytoplankton is typically limited by light and temperature, whereby the colder temperatures and lower light regimes are less conducive to algal growth. Conversely, during summer the warmer temperatures and higher light availability provides an environment more conducive to phytoplankton growth. During the summer months, phytoplankton production ultimately becomes limited by the amount of nutrients available for plant growth.

Climate also influences nutrient cycling, because significant quantities of nutrients may be carried into the Bay as a result of rainstorms (Sokolov and Black 1999), and some major toxic algal blooms have followed thunderstorms (Arnott et al. 1997). Sokolov and Black (1999) presented the concept of a dynamic pool of chemicals in the catchment, building up during dry periods, and running down in wet periods. Measurements from the Yarra River indicated that even a moderate storm event, after a dry period, leads to a substantial peak in riverine concentration of nutrients. Peaks from subsequent storms are increasingly damped. While the riverbed is an area of deposition (as indicated by the need for maintenance dredging), most of the nutrients carried in the water column are expected to be flushed into Hobsons Bay. An investigation of the environmental factors that influence the onset, development and timing of R. cf. chunii blooms in Port Phillip Bay was undertaken by Nicholson et al. (1998). This work involved the collation and analysis of a large database of phytoplankton and environmental variables. Cells of R. cf. chunii were noted to usually appear in June and July, with blooms, when they occurred, commencing in early August and ending in September or early October. The major blooms of R. cf. chunii have developed initially in the Geelong Arm, although blooms have also developed independently in Corio Bay. From the Geelong Arm, blooms have spread clockwise around the Bay and taken about 2-4 weeks to reach sites in the north and east (Nicholson et al. 1998). Due to the strong interannual synchronisation of the blooms, Nicholson et al. (1998) suggested that strong seasonal cues were operating to trigger cyst germination and bloom development. They hypothesised that strong winds would be required to disturb the bottom sediments and resuspend the algal cysts into the water column. This would be more prevalent in the Geelong Arm of the Bay, which is relatively shallow. Cysts of R. cf. chunii resuspended in the water column most likely sink back to the bottom, with germination only occurring at that time of the year when the environmental conditions are favourable. Germination itself is not influenced by nutrient concentrations, as nutrient uptake and cell division does not occur until afterwards. High ambient nutrient concentrations are, however, critical for bloom growth.

SEES Head Technical report: Nutrient Cycling 15

The increase in nutrient loads discharged from the Western Treatment Plant during winter, and its proximity to the Geelong Arm, are further possible reasons for the R. cf. chunii blooms first developing there. In contrast, nutrient concentrations in the south east of the Bay are generally lower than those in the north and west (Longmore et al. 1996), and this may explain why R. cf. chunii blooms have not developed or spread to this area of the Bay. Wind patterns, water temperature and light availability varied markedly in winter between years, with no evidence that bloom years were consistently different to non-bloom years. Therefore, a combination of several factors acting in concert seemed necessary to provide the conditions required to trigger cyst germination and subsequent bloom development (Nicholson et al. 1998). Two consistent indicators observed during bloom years were decreased salinity and the similarity in the composition of the phytoplankton community 4-6 weeks prior to bloom initiation. Nicholson et al. (1998) could not explain why reduced salinity was associated with bloom development, but suggested either salinity per se or a growth factor introduced with freshwater run-off was a requirement. The phytoplankton community composition in the weeks before August of bloom years was quite similar, and different to that in non-bloom years. Diatom concentrations were very low in June and July of bloom years, but relatively more abundant in these months in other years. These observations serve to indicate that though some climatic cues are obvious, it can be difficult to identify those factors which instigate or prolong algal blooms.

4.05 Alternative nitrogen recycling processes

Despite recent identification of an alternative nitrogen recycling process, studies indicate that denitrification is the dominant process for converting bioavailable nitrogen into nitrogen gas in Port Phillip Bay. As described above, the conventional view of denitrification in the sediments of Port Phillip Bay requires the juxtaposition of aerobic zones in the sediment, which promote the oxidation of ammonium to nitrate, and anaerobic zones, which allow the reduction of nitrate to N 2 gas. In the past few years, increasing interest has been shown in a microbial process called anammox (anaerobic ammonium oxidation). First identified in sewage treatment processes, anammox has more recently been observed in marine sediments ranging from the Arctic to tropical and temperate environments (Devol, 2003; Rysgaard et al. 2004; Engstrom et al . 2005), where it has accounted for up to 70% of the N 2 flux formerly attributed to denitrification. Anammox is also a microbially-mediated process, in which ammonium is oxidised by nitrite to produce N 2. Aerobic conditions are not required, but an external source of nitrite (or nitrate) is required. In many northern hemisphere estuaries, high concentrations of nitrate are found in the water column. However, this is not the case in Port Phillip Bay, and Heggie et al . (1999) demonstrated that denitrification in Port Phillip Bay sediments depends exclusively on nitrate generated within the sediments. Further, Thamdrup and Dalsgaard (2002) and Engstrom et al . (2005) found that the anammox contribution to N 2 production, compared to denitrification, declined with rate of carbon mineralisation, as indicated by benthic chlorophyll concentration. At the benthic chlorophyll concentrations found in Port Phillip Bay (Edmunds et al. 2004) denitrification is likely to be the dominant process. Risgaard- Peterson et al. (2005) studied the impact of MPB on anammox, and found that high levels of oxidized N were required in the water column to promote anammox. Oxygen production and nitrate uptake by MPB inhibited anammox, but not denitrification. They concluded that anammox probably plays an insignificant role in estuaries or coastal waters (such as Port Phillip Bay) which periodically exhibit N limitation. Rysgaard et al . (2004) also found that anammox activity was linearly related to bottom-water nitrate concentrations, and concluded that anammox bacteria were too slow-growing to be favoured under conditions of low or fluctuating nitrate.

SEES Head Technical Report: Nutrient Cycling 16

4.06 Measuring nutrient cycling

Nutrient cycling can be inferred from sediment nutrient profiles, or more reliably from direct measurements of benthic nutrient fluxes. Both approaches indicate the same spatial patterns, but only direct measurements provide an accurate estimate of the size of the fluxes in situ . The nutrient species found in sediments give some indication of the processes active in the sediment. The presence, absence, relative concentrations and /or changes with depth of the various forms reflect the chemical, biological and physical processes which result in the transfer of nutrients between the sediment and water column. Sediment sample collection carried out during the Port Phillip Bay Environmental Study contributed most to our understanding in this area, though samples were generally restricted to the top 30-70 cm. Chemicals migrate from areas of high concentration to areas of lower concentration by a process of diffusion. Nutrient concentrations are generally higher in sediments than they are in the overlying water column (Nicholson et al. 1996). This means that nutrients are more likely to diffuse from the higher- nutrient sediments to the lower –nutrient water column. This is termed a ‘positive nutrient flux’. Nicholson et al. (1996) found that diffusive fluxes varied strongly with location and time, but almost all were positive, indicating the sediment was a net source of dissolved nutrients. Ammonium diffusive fluxes varied six-fold in time and were highest in Hobsons Bay on all three sampling occasions. The next highest fluxes were found just south of Hobsons Bay. The highest phosphate diffusive flux was also found in Hobsons Bay in January 1994, and likewise varied five-fold over time. These observations are thought to reflect a higher supply of plant matter to the sediments in Hobsons Bay than elsewhere. Silicate diffusive fluxes appeared to reflect sediment type more closely than did ammonium and phosphate fluxes. Silicate fluxes were 2-5 fold higher at the muddy sites in central and northern Port Phillip Bay than in the sandier western and surround zones. Oxidised nitrogen diffusive fluxes were all within two standard deviations of zero. All fluxes were lowest at a sandy site off Mordialloc, again presumably reflecting a lower deposition of fresh organic matter at this site compared to the others. However, estimates of diffusive flux from pore water profiles should be viewed with caution. Diffusive flux estimates depend strongly on the slope of the concentration gradient in the upper few millimetres of the sediment. Nicholson et al. (1996) took relatively few pore water samples, and the interfacial concentration gradient was not well defined. In addition, nutrient flux estimates from concentration gradients ignore advective processes (e.g. fauna in the sediment pumping water), and photosynthetic processes in the MPB, which may both substantially modify fluxes (Berelson et al. 1998; Sundback et al. 1991). The uncertainties associated with the methods discussed above can be overcome through the direct measurement of nutrient fluxes from the sediment. The approach used in Port Phillip Bay since 1994 involves the deployment of benthic chambers at each site, which trap a volume of water over a known of sediment (Fig 7). The chambers are carefully lowered to the sediment surface, with penetration into the sediment monitored by video. Chambers are left in place for up to 24 hours. During this incubation time, nutrients exchange between the sediment and overlying water, and either accumulate or are depleted in the chamber. Sediment/water fluxes are estimated from linear regression of concentration on time. Replicate chambers provide an estimate of small-scale spatial variability. Fluxes are estimated for oxygen, alkalinity, carbon dioxide, ammonium, nitrite, nitrate, N 2, phosphate, silicate, bio-irrigation rate, and denitrification. Phosphate and silicate are included because of their importance in supporting phytoplankton blooms, and in determining bloom composition. Denitrification efficiency is calculated as the proportion of dissolved inorganic N (the sum of N 2, ammonium, nitrite and nitrate) recycled as N 2. If all of the recycled N is lost as N 2, efficiency is 100%. If all recycled N is returned as bioavailable ammonium and/or oxidised N, efficiency is zero.

SEES Head Technical report: Nutrient Cycling 17

Sample holders DO sensor

Benthic chamber Pump, microprocessor

Video camera

Figure 7 An automated benthic chamber used to directly measure nutrient fluxes between sediment and water column.

In situ measurements of benthic fluxes were first made with manual and automated benthic chambers at a number of sites in Port Phillip Bay during 1994-5. Since the first direct measurements in 1994 there have been more than 450 chamber incubations carried out in Port Phillip Bay. This includes 145 for the Port Phillip Bay Environmental Study (Nicholson et al. 1996; Berelson et al. 1998), 30 for the CSIRO Centre for Research into Introduced Marine Pests (Longmore et al . 1996a), 40 for the Port Phillip Bay Environmental Management Plan (Longmore 2005), 150 for an Australian Research Council Linkage project (Ross et al . unpublished) and 90 for the Channel Deepening Project (Longmore 2006b). These are summarised in Figure 8. In almost all of the measurements carried out since 2002, denitrification efficiency has been calculated from direct measurement of inorganic N and N 2 fluxes. Most benthic flux measurements carried out in Port Phillip Bay (Nicholson et al. 1996; Berelson et al . 1998) have demonstrated that the sediment is a net consumer of oxygen from the water column. Sediments with a MPB layer still showed net oxygen consumption over a full daily cycle.

SEES Head Technical Report: Nutrient Cycling 18

.64(5) 33(12) 96(2) . . 96(1) .81(5) . .76(12) .47(10) 96(5) .88(4) 69(2) . . .63(3) 73(2) 97(8) 82(7) . . 68(5) . . .48(48) 82(3) 85(3) . 91(4) 47(9) . . . 87(3) .78(12); 23(12) . .75(7)

.90(5) .90(1) .55(3)

.67(8) .95(5)

Figure 8 Denitrification efficiency measurements (mean and SE) for all studies since 1994. Black – PPBES 1994-96; blue ARC Linkage project 1992-94; green Channel Deepening Project 1994- present; red PPBES and PPB EMP 1994-present ( Berelson et al 1998; Longmore 2005; Ross et al. unpubl).

Nicholson et al. (1996) and Berelson et al . (1998) found that ammonium dominated the benthic dissolved inorganic nitrogen flux in Port Phillip Bay, but that there was much less nitrogen measured in the benthic chambers than expected from mass balance calculations. The difference was assumed to be in the form of nitrogen gas, from the coupled processes of nitrification and denitrification. This assumption was later verified by direct measurement of gas flux (Heggie et al. 1999) and, Bay-wide, about 52 % of primary production remineralised at the sediment is ultimately lost to denitrification (Murray and Parslow 1999). This is higher than that for many other coastal systems with comparable carbon loading (Berelson et al . 1998). Nicholson et al . (1996) estimated the Bay-wide benthic nitrogen flux from benthic chamber measurements: the dissolved inorganic nitrogen flux was about 3,600 tonnes N yr -1, and nitrogen gas flux about 15,000 tonnes N yr -1, implying a bay-wide denitrification efficiency of about 80%. This differs from the PPBES model estimates because of the differing treatment of the Sands area of Port Phillip Bay, and because the benthic flux measurements of Nicholson et al . were biased toward the deeper muddy sediments. The PPBES model assigned a zero efficiency to the Sands area, and included it in the Bay-wide average, whereas Nicholson et al. (1996) excluded the Sands from calculations.

SEES Head Technical report: Nutrient Cycling 19

Denitrification can also occur in the water column, but is significant only in very turbid waters (Omnes et al. 1996). The particulate and nitrate concentrations in Port Phillip Bay waters are too low for this to be a significant process (though it may be important in nitrate-rich turbid plumes from the Yarra River, or possibly in dredge plumes). In summary, indirect measurements of nutrient cycling indicated highest diffusive ammonium and phosphate fluxes in and south of Hobsons Bay, with up to five-fold changes in the size of the fluxes over 18 months. Direct measurements confirmed the spatial pattern, and indicated that the highly efficient denitrification over most of the Bay was significantly reduced in Hobsons Bay.

SEES Head Technical Report: Nutrient Cycling 20

5 Nutrient Processes in the Project Areas

This section discusses current understanding of significant features of nutrient cycling, inferred from water column measurements, sediment measurements and benthic flux measurements, in each of the Project Areas. The purpose of this approach is to provide the background information against which area-specific risks may be identified.

SEES Head Technical report: Nutrient Cycling 21

5.1 Yarra River and Hobsons Bay

The Project Area here extends from Swanston Dock to the outer line of the Williamstown breakwater (Fig 9). It includes the Yarra River channel, and the Williamstown Channel.

Figure 9 Yarra River and Hobsons Bay project area.

SEES Head Technical Report: Nutrient Cycling 22

The Yarra River is the main source of fresh water to Hobsons Bay, which is up to 11 m deep in the centre. The following general points can be made in relation to nutrient cycling within the water column and sediment of this project area. The Yarra River plume passes down the east coast, and under normal flow, most of the nutrients from the Yarra River are taken up by plankton in the river and Hobsons Bay. Light and temperature limit plankton growth in winter, while nutrients are limiting in summer. Response of the plankton to high-flow events occurs over days to weeks. Near-shore areas have an extended period of high phytoplankton productivity over spring, summer and autumn whilst further offshore, productivity stays low until a peak in summer, declining again in autumn. Toxic algae have bloomed in Hobsons Bay in summer, and the cysts of toxic algae have been found in Hobsons Bay, and in lower numbers in the Yarra River. Plankton growth-promoting nutrients likely to be released during dredging are highest in the surface metre of sediment from the Yarra River, and below 2 m in Williamstown Channel. As much may be released from particles over 10 days after dredging as is released during dredging. Benthic nutrient fluxes are currently higher in Hobsons Bay in summer than winter, while the opposite is true for denitrification efficiency: the usually low denitrification efficiency in Hobsons Bay is evidence of a system under stress.

Water column processes

Water circulation in the Yarra River estuary is dominated by seaward flow of freshwater in an upper layer, and a saline counter-current travels upstream on the bottom (Sokolov 1997). Circulation in Hobsons Bay is principally wind-driven; Grant and Gayler (1982) observed the Yarra plume heading either down the east coast, or toward the west, depending on the wind direction. There are regular measurements of water quality in the Yarra and Maribyrnong Rivers, but none are made in the estuarine areas of interest here. Though Sokolov (1997) indicated bi-directional flow in the Yarra estuary, with seaward flow in the surface layer, and landward flow in the bottom layer, the subsequent estimates of inputs into Port Phillip Bay did not assume any nutrients were retained in the estuary. An extensive set of water quality data has been collected from Hobsons Bay. This includes sampling carried out for the EPA Fixed Sites network (1986-present), Fisheries Victoria from 1990-97, and the PPB Environmental Management Plan (2002-present). Longmore et al. (1996) measured water quality at six sites around Port Phillip Bay at fortnightly intervals from 1990 to 1997 (Fig 10), including a site at Williamstown. Dissolved organic N dominates the N species at Williamstown, with occasional large oxidised N peaks (Fig 11). Peak ammonium and dissolved organic N concentrations at Williamstown occur in summer of most years, but there were also peaks in autumn 1990 and winter 1995. Ammonium (and oxidised N) are immediately available for plant growth, while dissolved and particulate organic N compounds have a range of reactivities, and may require microbial transformations before they become bioavailable (if at all). Grant and Gayler (1982), observed a series of intense short-lived blooms in Hobsons Bay from measurements at two-week intervals over the period 1974-76, with chlorophyll levels exceeding 30 µg L - 1(SEPP objective: median 2.5 µg L , 90 th percentile 4 µg L -1). These blooms were dominated by flagellates, rather than diatoms that otherwise dominated for the rest of the time. The blooms were restricted to Hobsons Bay. Highest productivity was at the Yarra mouth, declining further into Hobsons Bay. Blooms were restricted to the top 3 m of the water column. The chlorophyll a and primary production measurements may be underestimates, missing the activity of nanoplankton which accounts for up to 30% of biomass in Port Phillip Bay (Beardall et al . 1997) and would have passed through the filters used. Grant and Gaylor (1982) observed a five-fold increase in primary production between winter and summer, but there was a disjunction between productivity and grazing. In winter, biomass was high at low productivity, presumably because of low grazing. Both rose in spring, until grazing caught up, and biomass declined in late spring. Productivity and biomass were high in summer, then both declined through autumn. Laboratory incubations of pure cultures at saturating light indicated a linear relationship between productivity and temperature. They believed their observations were consistent

SEES Head Technical report: Nutrient Cycling 23

with light and temperature limitation of plankton growth in the winter, and nutrient limitation in the summer.

Maribyrnong River

r ive R W Melbourne ra e ar rr Y ib ee R iv Long Reef er Williamstown Little River * Hobsons Bay ek re C c Western Sandringham lo al Treatment di or M Plant Werribee River

er * Long Reef15 m 20 m iv R on rs * Central * Patterson tRte Corio * Corio Bay Clifton Springs Pa Bay Wooley Reef

Great Sands 20 m * Dromana Blairgowrie 0 15 km Bass Strait

Figure 10 Water quality sites sampled for Fisheries Victoria (1990-97, squares) and EPA (1986- present, stars). Note that continuous data is now collected from Central, Hobsons Bay and Long Reef sites for the PPB Environmental Management Plan (Longmore et al . 1996).

SEES Head Technical Report: Nutrient Cycling 24

Williamstown N

M) 12 µ µ µ µ Ammonium 10 Oxidised N 8 Particulate N Organic N 6 Concentration ( Concentration

0 Jan-90 Jan-91 Jan-92 Jan-93 Jan-94 Jan-95 Jan-96 Jan-97 Jan-98 Date

Figure 11 Nitrogen species concentrations in Hobsons Bay, 1990-97 (Longmore et al . 1997).

EPA monitoring of chlorophyll at a site in Hobsons Bay has been carried out since 1986. Data from 1989 – 2005 (Fig 12) indicated that peaks in chlorophyll concentration may be observed throughout the year, but more often in summer than other seasons. Of the 14 occasions in the data set on which chlorophyll concentration in Hobsons Bay exceeded 4 µg L -1, four were in spring, eight in summer and two in autumn. Data from a more intensive sampling program (Fig 13) also indicated most peaks in summer, with the two largest, in 1992 and 1993, attributed to blooms of the toxic dinoflagellate Alexandrium catanella . These seasonal patterns agree with the chlorophyll observations (more intensive, but for only one year) of Beardall et al . (1997), that is, a summer maximum and winter minimum in Hobsons Bay.

SEES Head Technical report: Nutrient Cycling 25

Hobsons Bay

10 ) -1 8 gL µ µ µ µ ( a 6

4 Chlorophyll

0 Jan-89 Jan-91 Jan-93 Jan-95 Jan-97 Jan-99 Jan-01 Jan-03 Jan-05 Date

Figure 12 Chlorophyll a concentrations from the EPA fixed site in Hobsons Bay, 1989-2005.

Williamstown

20 ) -1 gL µ µ µ µ 15 ( a

10 Chlorophyll

0 Jan-90 Jan-91 Jan-92 Jan-93 Jan-94 Jan-95 Jan-96 Jan-97 Jan-98 Date

Figure 13 Chlorophyll a concentrations at Williamstown, 1990-97 (Longmore et al . 1997).

SEES Head Technical Report: Nutrient Cycling 26

Beardall et al. (1997) carried out monthly measurements of phytoplankton biomass and productivity at 10 sites in Port Phillip Bay over the period 1994-95. Most of the variation in biomass was for the microphytoplankton (> 20 µm) and less for the nanoplankton (2-20 µm); the picoplankton biomass (0.2-2 µm) varied little. Beardall et al . (1997) contrasted the different response to increased biomass at central and southern sites, compared to Hobsons Bay. At all sites studied by Beardall et al . (1997) except two, increasing chlorophyll concentrations correlated with an increasing contribution from microplankton, which are more readily sedimented from the water column than the smaller cells. Increasing chlorophyll at Hobsons Bay was not related to cell size, indicating that increasing production as a result of nutrient inputs includes persistent smaller cells, which may be transported further in to Port Phillip Bay, rather than deposited locally. The toxic dinoflagellate Alexandrium catenella was first observed in Port Phillip Bay in 1986. Blooms of the alga have occurred intermittently between December and May in northern parts of the Bay, including Hobsons Bay, raising concerns regarding the accumulation of potent biotoxins in local shellfish (Arnott et al. 1994). In temperate waters A. catenella and other dinoflagellates form dormant cysts when environmental conditions become unfavourable. These cysts sink to the bottom after formation, where they become incorporated into the sediment. Cysts can survive even in dark anoxic sediments for long periods of time, and then act as seed populations for fresh blooms of motile cells when conditions return to favourable (Arnott et al. 1994). Investigation of the distribution of A. catenella cysts in sediments along the shoreline in the Bay found cysts in the northern Bay from Brighton to Werribee, including Hobsons Bay (Sonneman 1992, as cited in Arnott et al. 1994). Further work focussing on the concentration and viability of cysts from the Yarra River, Hobsons Bay and the Port of Melbourne spoil ground was undertaken prior to scheduled dredging operations (Arnott et al. 1994). A. catenella cysts were found to be abundant in Hobsons Bay (322-835 cysts g-1) including Webb Dock (600 cysts g -1), but much less common in Yarra River sediments. A sharp decline was observed between Webb Dock and the Yarra River just downstream from the Westgate Bridge (59 cysts g -1), with concentrations continuing to decrease further upstream. A. catenella cysts were also comparatively scarce in northern Port Phillip Bay outside of Hobsons Bay. Cysts were only found at one of six PoM DMG sites at a concentration of 73 cysts g -1, lower than that of Hobsons Bay. These could have been transported by dredge material, although low concentrations of cysts were also found in control areas situated east and west of the PoM DMG, with one of three sites in each containing 23 cysts g-1 and 43 cysts g -1 respectively. The control site cysts may have been dispersed from the PoM DMG, or arisen through transportation of motile cells from Hobsons Bay during an earlier bloom (Arnott et al. 1994). Cysts of other non-toxic dinoflagellate species were abundant at most sites, including the PoM DMG and control sites, with concentrations again greatest in Hobsons Bay and lower in the Yarra River. This suggests the low concentrations of A. catenella cysts at the PoM DMG and control sites were unrelated to sediment grain size. Furthermore, these sediments had a grain size distribution with a high percentage of silt (<63 µm), usually suitable for cysts (Arnott et al. 1994). A. catenella cysts were successfully germinated only from sediments sampled from Hobsons Bay and the Yarra River just downstream from the Westgate Bridge, with motile cells identified from each. Cysts from the other sites did not germinate, but may still have been viable if given the period of dormancy and environmental conditions required to trigger germination (Arnott et al. 1994). Continuous sampling of dissolved oxygen concentration has been carried out in Hobsons Bay since 2002 (Fig 14). Measurements of dissolved oxygen concentration averaged 95-100% in near-surface waters, and 90-95% in bottom waters. Bottom DO concentration was often substantially below surface concentrations, indicating oxygen consumption exceeding supply. Conversely, there were occasions in which surface DO concentrations were super-saturated (>100%). Most, but not all, of the peaks in surface DO concentration coincided with peaks in chlorophyll concentration, indicating active production by plants (Longmore 2005). There were occasions when bottom DO concentrations also exceeded saturation, and in general these also coincided with peaks in bottom-water chlorophyll. On the other hand, peaks in bottom-water chlorophyll did not always lead to peaks in DO; often they were accompanied by declines. Processes, which resuspend the sediments (e.g. strong winds), may cause this pattern (high chlorophyll and low DO).

SEES Head Technical report: Nutrient Cycling 27

The indicators (salinity, temperature, dissolved oxygen and chlorophyll fluorescence) monitored continuously were also studied to determine the response to the February 2005 floods in the Yarra and Werribee Rivers (Longmore 2005). Chlorophyll a concentration increased for about 10 days after the flood in Hobsons Bay, with a greater increase in bottom waters than in surface waters. This could be either a response to the settling of riverine plant material, or resuspension of MPB from the sediment. A secondary peak was observed in surface waters 4-5 weeks after the flood. Dissolved oxygen concentrations responded immediately to the flood, with an increase to supersaturation in surface waters, and a decline to below 70% in bottom waters (Fig 14). Concentrations recovered to above 80% within a month. The diurnal range in surface DO concentrations reached 30% following the flood, and fell to less than 10% within a month.

150

140

130

120

110 10 m 3 m 100 DO (% sat) DO

60 24-May-02 10-Dec-02 28-Jun-03 14-Jan-04 1-Aug-04 17-Feb-05 5-Sep-05 Date

Figure 14 Continuous records of dissolved oxygen concentration from near-surface and near- bottom waters in Hobsons Bay, 2002-05 (Longmore 2005).

Benthic processes

Gorfine et al. (1996) measured NaCl-extracted ‘interstitial and loosely bound’ inorganic ammonium and phosphate in triplicate sediments from 40 sites in Port Phillip Bay in 1991. They found a bay-wide average of 12.2 μM ammonium and 8.35 μM phosphate in the top 2 cm of sediment. Highest concentrations were found in Hobsons Bay. They did not analyse deeper sediments, nor relate the extracted nutrients to those likely to exist naturally in pore waters. The only available sediment data collected from the Yarra River is that for the EES (SKM 2004) and the SEES (Longmore 2006c). The EES sediments were collected principally for toxicant studies, and provided little information relevant to nutrient cycling. Analyses for pore water ammonium concentration were made on a limited number of cores. The ammonium concentrations found in sediments from the Yarra River were generally above the range found by Nicholson et al. (1996) in Hobsons Bay sediments. One sediment core in particular produced pore water ammonium concentrations of more than 1,800 µM to 1.2

SEES Head Technical Report: Nutrient Cycling 28

m depth, some four times greater than the highest concentrations measured by Nicholson et al. (1996). Ammonium was usually richer (average of 12-23 mg N kg -1) in the surface sediments in each area than in the sub-surface sediments (7-18 mg N kg -1). However, in the sediments between the Westgate Terminal and the Westgate Bridge, the sub-surface layer was much more enriched (60 mg N kg -1 ) than the surface layer (16 mg N kg -1). Longmore (2006c) measured dissolved inorganic nutrients elutriated from sediments in cores collected from seven sites in the Yarra River and four sites in the Williamstown Channel. Ammonium and oxidised N concentrations in the Yarra River were about twice as high in elutriates from the surface metre of sediment than at greater depths, while phosphate and dissolved organic N concentrations were 10-20 times higher in the surface metre, and silicate concentration varied little with depth. The loads of bioavailable nutrients likely to be released in the hopper or in overflow have been calculated from these measurements and estimates of in situ sediment volumes to be dredged, as has an estimate of ammonium-N load from the earlier SKM work (Table 2).

Table 2 Estimates of dissolved nutrient load (median ±±± 95% CI, tonnes ) in Yarra River sediments to be dredged (SKM 2004; Longmore 2006c).

Ammonium Oxidised N DON Phosphate Silicate 111*

16.4 ± 23.9 0.03 ± 0.02 10.0 ± 11.1 1.0 ± 0.7 19.8 ± 5.5 *From SKM (2004)

The most recent estimate is thought to be more accurate than the earlier one (SKM 2004), because it is drawn from a greater number of cores, and a greater range of depths.

The most comprehensive survey of pore water nutrient concentrations in Hobsons Bay was carried out by Nicholson et al. (1996), as part of the PPBES. Nicholson et al. (1996) collected triplicate sediment cores from two sites in July 1993, and resampled one site (outside the shipping lanes) on two more occasions over the next year (1993-94), collecting pore waters to 20-70 cm. Small-scale spatial and temporal variability was significant. Typical profiles from Hobsons Bay indicated highly heterogeneous sediments, with pore water profiles from one replicate core often quite different to the other two replicates, even though cores were collected within a distance of 1 m of each other (Fig 15). Even so, general patterns emerge of an increase in nutrient concentration with depth to 5-10 cm, then a nearly constant concentration, or slight decline, at greater depths. Pore water dissolved organic N concentrations (not shown) were similar to ammonium concentrations, while dissolved organic P concentrations were only about 10% of the phosphate concentrations (Nicholson et al . 1996). Bay-wide, highest pore water inorganic nutrient concentrations were observed in Hobsons Bay and the Yarra mouth. The cores indicated a much higher ammonium concentration in the sediment than in overlying waters, suggesting that there should be a diffusion-driven flux from the sediment. Estimates of diffusive ammonium fluxes varied six-fold in time (0.2-1.3 m mol N m -2.d -1) and were highest in Hobsons Bay on all three sampling occasions (Nicholson et al. 1996). Phosphate diffusive flux estimates likewise varied five-fold at the Hobsons Bay site over time. Oxidised nitrogen diffusive flux estimates were all within two standard deviations of zero. SKM (2004) also measured ammonium concentrations in some cores from the Williamstown Channel. Longmore (2006c) measured dissolved inorganic nutrients elutriated from sediments in cores collected from four sites in the Williamstown Channel. Ammonium, phosphate and silicate concentrations were 4- 30 times higher in elutriates from 1-4 m deep than in the surface metre. The means are highly biased by much higher concentrations from a core near the Williamstown dockyards, than in the other three cores.

SEES Head Technical report: Nutrient Cycling 29

30 July 1993; Hobsons Bay

Ammonium (uM) 0 20 40 60 80 100 120 140 0

15 Core 1 Core 2 20 Core 3 Depth (cm) Depth

Figure 15 Small-scale spatial variability in sediment pore water ammonium concentration: three cores collected 1 m apart in Hobsons Bay (Nicholson et al. 1996).

The highest concentrations of elutriated ammonium (almost 2,500 µM) were found at three depths between 1-4 m in one core (VB139) opposite the Williamstown dockyard. All other ammonium concentrations from the Williamstown Channel were less than 1,000 µM. The loads of bioavailable nutrients likely to be released in the hopper or in overflow have been calculated from these measurements and estimates of the in situ volume of sediment to be dredged, as has an estimate of ammonium-N load from the earlier SKM work (Table 3).

Table 3 Estimates of bioavailable nutrient load (tonnes, median ±±± 95% CI) in sediment to be dredged from the Williamstown Channel (SKM 2004; Longmore 2006c).

Ammonium Oxidised N DON Phosphate Silicate 19.5*

17.7 ± 71.0 0.01 ± 0.01 5.8 ± 9.0 0.8 ± 1.5 5.6 ± 3.7 *From SKM (2004).

The estimated bioavailable ammonium and phosphate loads in the Williamstown Channel are similar to that in the Yarra River, despite the larger volume of sediment to be dredged in the Yarra River. The other nutrient loads are proportional to the dredged volume. When sediment is dredged by suction, some bioavailable nutrients are released into the suction water, and are discharged in the overflow. In addition, nutrients may be converted from particle-bound to

SEES Head Technical Report: Nutrient Cycling 30

bioavailable forms over time, as sediment particles settle from the dredge plume. New data on sediment N, P and biogenic Si content, and longer-term release of bioavailable nutrients, has been generated on 11 cores collected from the Yarra River and Williamstown Channel (Longmore 2006c). There was no strong pattern of total N concentration with depth in the Yarra/Williamstown Channel cores, but the three highest values (to 2,200 µg N g -1) occurred in the surface 30 cm. Sediment total P concentrations showed a similar pattern to the sediment N concentrations, that is, near-surface enrichment, but no other consistent pattern with depth. .Biogenic Si concentrations in the Yarra-Williamstown Channel area showed a wide range (<1 to 7.5 mg Si g -1), with no overall trend with depth. When multiplied by the expected in situ volume of sediment to be dredged, substantial loads of nutrient will be relocated by dredging in the Yarra River and Williamstown Channel (Table 4).

Table 4 Total nutrient load (tonnes, mean ±±± 95%ile) in sediments to be dredged from the Yarra River and Williamstown Channel (Longmore 2006c). Nutrient Yarra load Williamstown load Total N 2,200 ± 6,250 1,050 ± 1,450 Total P 975 ± 600 460 ± 380 Biogenic Si 16,900 ± 2,900 7,200 ± 5,600

When sediment samples were suspended in seawater and incubated for 10 days, nutrient release from the sediment was observed (Longmore 2006c). Highest concentrations, and greatest increases over time, were found at depth in core VB139, in the Williamstown Channel. Nutrient release rates slowed over time for some samples (eg. all those from the Yarra River), while others increased at an undiminished rate over the 10 days of the experiment. The release rate was related to the volume of sediment to be dredged to calculate a load (Table 5). These calculations suggest that as much ammonium may be released from suspended or settled particles over 10 days as is released immediately from the pore water during dredging.

Table 5 Mean nutrient release over 10 days (tonnes ±±± 95% CI): change in individual samples averaged and multiplied by dredge volume for each project area and depth interval (Longmore 2006c). Area Ammonium Oxidised N DON Phosphate Silicate Yarra River 20.8 ± 14.4 0.7 ± 0.7 -1.8 ± 3.0 1.3 ± 1.8 14.8 ± 5.5 Williamstown 23.8 ± 20.0 0.2 ± 0.2 1.0 ± 2.3 0.6 ± 0.9 7.7 ± 3.6 Channel

No direct measurements of nutrient cycling have been made in the Yarra River. The PPBES model (Murray and Parslow 1997) predicted low (<50%, and sometimes zero) denitrification efficiency in the Yarra River. A number of measurements of nutrient fluxes have been made in Hobsons Bay, with most at the one site occupied for the PPBES, Port Phillip Bay Environmental Management Plan and CDP baseline monitoring. The measurements indicate consumption of oxygen, and production of carbon dioxide, ammonium, oxidised N, phosphate, silicate and N 2 (Table 6). Mean oxygen flux exceeds carbon dioxide flux, indicating some non-metabolic oxygen consumption (eg. via oxidation of reduced sulphur compounds). There is also evidence from individual incubations (Berelson et al . 1998) that a significant proportion of the organic matter in Hobsons Bay is oxidized by sulphate, rather than oxygen. The nitrogen flux is dominated by N 2, closely followed by ammonium. Oxidised N fluxes are much smaller than ammonium

SEES Head Technical report: Nutrient Cycling 31

fluxes. Because of the similarity between DIN (ammonium + nitrite + nitrate) and N 2 fluxes, denitrification is about 50% efficient: that is, about half of the recycled inorganic nitrogen is lost as N 2.

Table 6 Summary of direct benthic flux measurements ( m mol m -2 d -1) in Hobsons Bay, 1994- 2005 (Nicholson et al. 1996; Berelson et al 1998; Longmore 2005). Variable Flux, mean ±±± 1 SE. Dissolved oxygen -42 ± 2.6 Carbon dioxide 36 ± 4.2 Ammonium 2.1 ± 0.4 Oxidised N 0.2 ± 0.03 Phosphate 0.30 ± 0.05 Silicate 7.1 ± 1.1

N2 3.1 ± 0.6 Denitrification 56 ± 4.7 efficiency (%)

At the Hobsons Bay site, all fluxes, excluding dissolved oxygen, ammonium and denitrification efficiency, varied significantly between years. The data indicate generally higher fluxes in summer and autumn than in spring, and reflect the seasonal variation in supply of organic matter to the sediments in Hobsons Bay (Fig 16). In Hobsons Bay in spring 2002, denitrification efficiency was higher than the historic median, which could be attributed to a lower than normal organic supply, due to low riverine inputs (Fig 17). However, carbon dioxide fluxes in later samples were closer to previous measurements, as was denitrification efficiency. There is some evidence of a fall in denitrification efficiency in the autumns of 2003, 2005 and 2006, compared to the previous spring (Fig 17). Because this fall has occurred often enough to be a general feature, the change cannot be attributed to “patchiness” at the sample sites. Examination of the dilution rate of a tracer added to the chambers, as it is mixed into the porewater, shows appreciable change from one period to the next (Longmore, unpublished). However, the changes are not uniform in either magnitude or direction from spring to the next autumn of each of the recent years, which appears to rule out changes in bio-irrigation as a potential explanation of the seasonal differences. Nor is it directly related to changes in the amount of organic matter being degraded: the change in carbon dioxide flux does not always match the change in denitrification efficiency. Nor does it appear to be evidence of a long-term decline in bay health. These seasonal differences have occurred during a prolonged drought in the Bay catchment. On an annual basis, none of the differences in nutrient fluxes were consistent over sampling periods (ie. there were no trends over the whole data set). Rather, the differences were usually due to a specific year. For example, phosphate and silicate fluxes from 1994 were statistically significantly different (p<0.05) to all subsequent measurements, while carbon dioxide fluxes in 1995 were significantly different to fluxes in 1996 and 2003-05. As discussed above, denitrification efficiency was lower in autumn 2003, 2004 and 2005 than in the previous spring, but the differences between years were not statistically significant (Longmore 2005).

SEES Head Technical Report: Nutrient Cycling 32

Carbon dioxide flux, Hobsons Bay

100

80 ) -1 70 d -2 60

30 flux (m mol C(m mol flux m 2 20 TCO 10

-10 Sp Su Wi Su Sp Au Sp Au Sp Au Sp Au 1994 1995 1995 1996 2002 2003 2003 2004 2004 2005 2005 2006

Figure 16 Variation in carbon dioxide flux (mean ±±± 1 SD) in Hobsons Bay, 1994-2006 (Nicholson et al. 1996; Berelson et al 1998; Longmore 2005).

Denitrification Efficiency, Hobsons Bay

120

100

40 Denitrification Efficiency (%) Efficiency Denitrification 20

0 Sp Su Wi Su Sp Au Sp Au Sp Au Sp Au 1994 1995 1995 1996 2002 2003 2003 2004 2004 2005 2005 2006

Figure 17 Variation in denitrification efficiency (mean ±±± 1 SD) in Hobsons Bay, 1994-2006 (Nicholson et al. 1996; Berelson et al 1998; Longmore 2005).

SEES Head Technical report: Nutrient Cycling 33

Summary for the Yarra River and Hobsons Bay

Key nutrient cycling processes : • Light and temperature limit plankton growth in winter, and nutrients are limiting in summer; • Phytoplankton biomass and productivity is highest in summer; • The potential for toxic algal blooms is highest in summer; • Benthic nutrient fluxes in Hobsons Bay are higher, and denitrification efficiency lower, in summer than in winter. • The Yarra River sediments are not thought to have a significant capacity to denitrify; Hobsons Bay sediments have a denitrification efficiency of about 50%, indicative of a system under stress.

Key factors affecting processes:

• Nitrogen in the water column is generally highest in summer in Hobsons Bay; • During high flow, much of the nutrients delivered by the Yarra pass through Hobsons Bay and impact further south; under base flow, most of the nutrients are assimilated within Hobsons Bay; • The Yarra plume follows the east coast under prevailing westerly winds, but may move toward Altona under north-easterlies; • Because of small cell size, organic matter associated with algal blooms in Hobsons Bay may be transported further into the northern part of Port Phillip Bay; • The system responds to natural events (eg. a flood in the Yarra) by a rapid increase in plankton biomass over about 10 days; a second peak some 4-5 weeks later may indicate the lag between algal bloom deposition and regeneration from the sediment; • The concentration of bioavailable nutrients likely to be mobilised in a dredge hopper is higher in the surface metre of sediment than deeper in the Yarra River; the opposite is true for the Williamstown Channel; • The load of bioavailable nutrients likely to be mobilised in a dredge hopper is higher in the Williamstown Channel than the Yarra River; • A similar load of nutrients may become bioavailable over 10 days after suspension in the dredge plume.

SEES Head Technical Report: Nutrient Cycling 34

5.2 North of the Bay

The area of the Bay under consideration here extends from south of Williamstown in the north, to a line between Clifton Springs and Mornington (Fig 18). It includes the waters north of the Sands, the central area of the Bay up to 25 m deep, and the shallow western waters, which receive the largest single nutrient discharge to the Bay, from the Western Treatment Plant.

Figure 18 Northern project area. Strong nutrient gradients exist in the water column in this area, especially near the coasts, Werribee and Hobsons Bay, but the impact of inputs may be widespread: the February 2005 Yarra River flood led to a response in the centre. The timing of peak phytoplankton biomass varies from winter in the centre and the west, to summer in the north. Port Melbourne Channel sediments are nutrient-poor, compared to those from Hobsons Bay, and the ammonium load in the sediments to be dredged is only about 10% of previous estimates. Benthic nutrient fluxes in central Port Phillip Bay are much lower, and denitrification efficiency much higher (at ~ 85%), than in Hobsons Bay. There are significant differences in fluxes between years at the one site, and between sites in the area, with lower efficiency often observed closer to shore. Because of the high efficiency, and large area, this area contributes significantly to the capacity of the Bay to denitrify.

SEES Head Technical report: Nutrient Cycling 35

Water column processes

Walker (1999) estimated the flushing time of water in northern Port Phillip Bay (essentially anywhere north of the Sands) at 270-280 days. Circulation is dominated by the wind (Fig 5), which generates a southward drift of the Western Treatment Plant discharge in winter, and directs the Yarra River plume along the eastern coast. Water movement through the Bay centre is weak, and arises from the interaction of the gyres that form in different locations under different winds. The EES “North of the Bay” area includes a number of sites sampled by a range of programs. These include the central site sampled by EPA since 1986 and by Longmore et al. (1996) between 1993 and 1995; the Long Reef site sampled by EPA since 1995 and Longmore (2005) since 2002, and two coastal sites (Sandringham and Werribee) sampled by Longmore et al. (1996) between 1990 and 1996. There is a strong gradient in nutrient concentrations from the northern near-shore waters to the centre of the Bay, and from the Werribee region toward the centre (Fig 19). Even so, some features are common to both areas, including a dominance of the nitrogen concentrations by dissolved organic nitrogen (Fig 20). The relationship between nutrient input and plankton growth is often difficult to observe in the field.

Ammonium (uM) Phosphate (uM)

14.0 5.0 12.0 4.0 10.0

8.0 3.0

6.0 2.0 4.0 1.5 2.0

1.5 1.0

1.0 0.5 0.5 0.0 0.0

Figure 19 Surface water nutrient concentrations, interpolated from 12 monthly underway surveys in 1993 (derived from Shao and Fox 1996).

SEES Head Technical Report: Nutrient Cycling 36

Sandringham

20 M) µ µ µ µ 15 Ammonium Oxidised N Organic N 10 Particulate N Concentration(

0 Jan-90 Jan-91 Jan-92 Jan-93 Jan-94 Jan-95 Jan-96 Jan-97 Jan-98 Date

Figure 20 Nitrogen species concentrations in northern Port Phillip Bay, 1990-97 (Longmore et al . 1997).

Seasonal peaks in dissolved inorganic nitrogen concentration reflect the importance of the Western Treatment Plant discharge to the western side of the Bay, while summer rains led to high concentrations of inorganic N in the north. The Western Treatment Plant was the only significant surface source of phosphate. Nutrients from the Western Treatment Plant are discharged to the western side of Port Phillip Bay, principally as dissolved inorganic nitrogen (ammonium and nitrate) and dissolved inorganic phosphate. There is a large seasonal variation in ammonium discharge from the Western Treatment Plant, from less than 2 tonnes N d -1 in summer to more than 11 tonnes N d -1 in winter (Murray 1994). Phosphate discharge is also greater in winter than summer (about 3 tonnes P d -1 and 1 tonne P d -1 respectively). Planktonic biomass was highest off the Western Treatment Plant in winter, and off Williamstown in summer (following peak river flow). It can be shown that most, if not all, of the dissolved inorganic nitrogen in the Western Treatment Plant discharge is stripped from the water column close to the discharge (Longmore et al . 1996). Variations in dissolved inorganic nitrogen background at other sites must then be due to recycling of organic matter both within the water column (“sloppy feeding” by zooplankton) and at the sediment surface (bacterial consumption producing benthic fluxes), rather than external inputs. Annual loads are available from the WTP to 2005 (Fig 21), but monitoring of riverine inputs is acknowledged to have been too sporadic until 2002 to generate accurate loads to the Bay (Sokolov and Black 1999; Parslow et al . 1999). Proposals are now under way to allow Melbourne Water to better measure loads to the Bay from the rivers (Fletcher and Deletic 2006). WTP loads varied between about 3,000 and 4,000 tonnes N per year between 1984 and 2001. The decline in N load since 2001 is attributed to the commissioning of an activated sludge plant, and reduced inflows in dry years.

SEES Head Technical report: Nutrient Cycling 37

4500

4000

3500

3000 Annualload(tonnes N)

2500

2000 1980 1985 1990 1995 2000 2005 2010 Year ending June.

Figure 21 Annual N load to the Bay from the Western Treatment Plant (Trevor Gulovsen, Melbourne Water, and Melbourne Water Annual Social and Environmental Reports).

Particularly high runoff in September 1993, while it did not lead to a large spike in ambient nutrient concentrations at Sandringham (Fig 20), did lead to a prolonged (~ 2 month) phytoplankton bloom in the north, producing the highest chlorophyll concentrations at Sandringham in 16 years. The following year, 1994, was dry, and only a small bloom was observed at Sandringham. There was a steady decline in mean bay-wide chlorophyll concentration through the period May 1993- January 1995, which was interpreted as a response to reduced inputs from the catchment (Murray and Parslow 1997). At the central bay site, of the five occasions on which EPA measurements of chlorophyll concentration exceeded 2 µg L -1, one was in summer (Jan 93), two were in autumn (May 90 and Mar 99) and two in winter (June 89 and Jul 01) (Fig 22).

SEES Head Technical Report: Nutrient Cycling 38

EPA Central site

2.5 ) -1 2 gL µ µ µ µ ( a 1.5

1 Chlorophyll

0.5

0 Jan-89 Jan-91 Jan-93 Jan-95 Jan-97 Jan-99 Jan-01 Jan-03 Jan-05 Date

Figure 22 Chlorophyll a concentrations from the EPA Central site, 1989-2005. Continuous dissolved oxygen and chlorophyll a measurements have been carried out at the 25 m deep central site since 2002 for the PPB EMP (Longmore 2005). These indicate much the same features as similar measurements in Hobsons Bay. Dissolved oxygen concentration averaged 95-100% in near-surface waters, and 90-95% in bottom waters. Bottom DO concentration was often substantially below surface concentrations, indicating oxygen consumption exceeding supply. At the Central site, in response to the February 2005 flood, surface chlorophyll concentration increased initially to ~ 3 µg L -1 at the surface, and gradually declined over the next six weeks, while bottom concentrations gradually increased, and remained above 1.5 µg L -1 until May 2005 (Fig 23). A second, much larger response (to 6 µg L -1 ) was observed in surface waters six weeks after the flood, but persisted for only three days. Dissolved oxygen concentration responded immediately to the increased biomass following the flood, leading to supersaturation in surface waters. The diurnal range also increased at the surface to ~ 10%. There was no diurnal signal at the bottom, but concentration fell to less than 80% for about a week following the flood, then gradually recovered over the next month to close to surface concentrations. Beardall et al . (1997) demonstrated that algal blooms in Port Phillip Bay outside of Hobsons Bay are dominated by large, rapidly sinking cells (diatoms). The area of impact of blooms on benthic nutrient cycling is therefore expected to coincide with the location of the blooms. Coastal sites at Beaumaris, Sandringham and Werribee sampled by Arnott et al . (1997) were also dominated by diatoms. Biomass peaked in winter and spring, while biomass in the centre of the Bay was highest in early winter (this observation was also supported by the continuous monitoring carried out in the centre of the Bay by Longmore 2005, who found lowest biomass in spring).

SEES Head Technical report: Nutrient Cycling 39

5 ) -1 4 g L µ µ µ µ (

a 17 m 3 3 m

2 Chlorophyll

0 11-Feb- 16-Feb- 21-Feb- 26-Feb- 3-Mar- 8-Mar- 13-Mar- 18-Mar- 23-Mar- 28-Mar- 05 05 05 05 05 05 05 05 05 05 Date

Figure 23 Chlorophyll measurements at the Central site, early 2005: Yarra flood response (Longmore 2005).

Benthic processes

Cores were collected from the sediments in the north of the Bay from seven sites, including four in the deep muddy sediments, one in sandy eastern sediments, and two in sandy sediments off Werribee (Nicholson et al . 1996), and were subjected to detailed nutrient analysis. None of the cores were from shipping channels. Cores were also collected from the Port Melbourne Channel for the EES (SKM 2004), with limited analyses of interest here. Sediment slurry samples were collected from the hopper of the Queen of the Netherlands while carrying out trial dredging in the Port Melbourne Channel (Longmore 2005b), and these samples were analysed for ammonium, oxidised N and total N concentrations. Further core collection has recently been carried out from the Port Melbourne Channel and PoM DMG for estimation of bioavailable nutrients released in the hopper or overflow (Longmore 2006c). The sediment nutrient analyses of Nicholson et al. (1996) revealed similar patterns with depth at all the muddy sites, that is, increasing concentrations of ammonium, phosphate, oxidised N and silicate with depth to about 5 cm, then either a decline or constant concentration below 5 cm. In contrast, solid-phase N, C and P (to a lesser extent) declined with depth, indicating that the most labile material is at the sediment surface. Pore water nutrient concentrations were not very different for the sandy sediments from near the Patterson River (Nicholson et al. 1996), but the total N, P and C concentrations were lower than for the muddy sediments from deeper waters.

SKM (2004) found dissolved organic carbon concentrations of 4-5 mg L -1 in PoM DMG (7 samples). SKM (2004) found that total organic carbon in deeper (2.9-3.0 m) sediments was very low (<0.2 %) for all cores from the Port of Melbourne Channel. On the basis of a limited number of analyses for ammonium in Port Melbourne Channel sediments by SKM, and anticipated volume of sediment to be dredged in the channel, an estimate was made of ammonium load in the dredged sediment (Table 7). This is combined with estimates from the most recent coring. Longmore (2006) found that the concentrations of all elutriated nutrients in eight cores from the Port Melbourne Channel were much lower than those from

SEES Head Technical Report: Nutrient Cycling 40

the Yarra and Williamstown Channel. DON was much higher in the surface metre than deeper, while the other nutrients changed little with depth. In contrast, elutriated nutrient concentrations increased steadily with depth for the five cores from the extension to the Port of Melbourne PoM DMG, and concentrations approached those found for the Yarra River. It is possible that the sediments sampled south of the existing PoM DMG are remnants of earlier disposal of Yarra River sediment. The estimates of ammonium release from the Port Melbourne Channel sediments (based on these analyses and the estimated in situ volume of sediment to be dredged) are only about 10% of that estimated for the EES.

Table 7 Estimates of dissolved nutrient load (tonnes, mean ±±± 95% CI) in sediment to be dredged from the Williamstown Channel (SKM 2004; Longmore 2006c).

Area Ammonium Oxidised N DON Phosphate Silicate Port Melbourne 26.3* Channel Port Melbourne 3.1 ± 2.8 0.02 ± 0.01 4.3 ± 3.4 0.9 ± 0.6 7.0 ± 2.5 Channel *From SKM (2004)

The total N loads calculated for the sediments removed in the TDP (44 ± 4 t N), when extrapolated to the in situ volume expected to be dredged in the same areas for the capital dredging program, were only 4% of those predicted for the EES. The most recent measurements (Table 8) confirm that the total N load to be mobilised during dredging in the Port Melbourne Channel (970 t) is less than 20% of that estimated for the EES ( 6150 t; Longmore 2004).

Table 8 Estimates of total nutrient load (tonnes ±±± 95% CI) to be mobilised during dredging in the Port Melbourne Channel (Longmore 2006c).

Port Melb Ch In situ volume dredged 2.55 (million m 3) Mean N load 740 ± 310 Mean P load 455 ± 240 Mean Si load 8,300 ± 2,200

Dredging necessarily leads to the resuspension of sediment particles. The nutrients, and specifically nitrogen, in sediments may be released into the water column, and thus stimulate plant growth. How far the nutrients are dispersed, and how effective they are at stimulating plant growth are complex questions, but in principle the nutrients associated with the finest particles will remain suspended for longer than coarser particles, and should therefore disperse further. No measurement has previously been made of the distribution of nitrogen with particle size in Port Phillip Bay sediments. Sediments were collected from the hopper of a dredge during the Trial Dredge Program (TDP), while dredging in the Port Melbourne Channel. Ammonium, oxidised N and dissolved organic nitrogen (DON) concentrations were measured in the aqueous phase. Wet sediment was suspended in a column of seawater, and particles of sizes 2-63 µm sampled. These were analysed for total nitrogen concentration, to determine the distribution of N across the particle size range. Sediments from the Port Melbourne Channel varied from fine to coarse, with up to 40% of the weight associated with particles less than 63 µm in diameter (silts and clays; Longmore 2006a). Ammonium and

SEES Head Technical report: Nutrient Cycling 41

DON concentrations in the aqueous phase of samples from the Port Melbourne Channel were similar to those measured in a dredge hopper during maintenance dredging in 2003. Nitrate formed only a minor part of the total dissolved N concentration. The dissolved N loads estimated to have arisen from the TDP of 1.6 ± 0.3 tonnes N in the north were 33% of those predicted in the EES (scaled to the volume dredged in the TDP). This is attributed to the channel sediments being organically-poor compared to those from Hobsons Bay on which the EES estimates were based. Those particles carrying the greatest proportion of the N load in the channel sediments were in the range 2-16 µm; such particles may remain suspended for more than 8 hours before they reach the seafloor. The potential therefore exists for some fraction of the particle-bound N to become bio-available, and experimental work has recently been completed to resolve this issue. Surficial sediments were the main source of bioavailable ammonium and phosphate released over 10 days in the Port Melbourne Channel (Table 9). The fact that ammonium and/or oxidised N increases did not relate to decreases in DON suggests that the bioavailable N is released from particles, rather than breakdown of DON. The most interesting fact from this work is that almost all the Port Melbourne Channel ammonium is released from particles over time, rather than directly in the pore water. The bioavailable ammonium release from particles is similar to the estimate provided to the EES.

Table 9 Estimates of bioavailable nutrient load released over 10 days (tonnes ±±± 95% CI) in sediment to be dredged from the Williamstown Channel (Longmore 2004; Longmore 2006c).

Area Ammonium Oxidised N DON Phosphate Silicate Port Melbourne 3-148 (32) range and most likely estimate provided to EES Panel Channel Port Melbourne 31.5 ± 40.8 1.8 ± 2.4 1.2 ± 2.3 2.0 ± 2.0 20.1 ± 19.1 Channel

Direct flux measurements are available from seven sites in northern Port Phillip Bay from the PPBES (Nicholson et al . 1996); eight sites for the ARC Linkage project (Ross et al . unpublished), one site for the PPB EMP (Longmore 2005), and three sites for the CDP baseline and TDP studies (Longmore 2006b). The data consistently indicate consumption of oxygen, and production of carbon dioxide, ammonium, phosphate, silicate and N 2, but all fluxes are much lower than those measured in Hobsons Bay (Table 10). Higher fluxes have been observed in autumn than in spring. The key indicator, denitrification efficiency, has always been higher in central Port Phillip Bay than in Hobsons Bay.

Table 10 Summary of direct benthic flux measurements ( m mol m -2 d -1) in central Port Phillip Bay, 1994-2005 (Nicholson et al. 1996; Berelson et al . 1998; Longmore 2005). Variable Flux, mean ±±± 1 SE. Dissolved oxygen -20 ± 1.1 Carbon dioxide 16 ± 1.1 Ammonium 0.26 ± 0.04 Oxidised N 0.14 ± 0.04 Phosphate 0.07 ± 0.01 Silicate 2.2 ± 0.2

N2 1.8 ± 0.3 Denitrification 85 ± 3.0 efficiency (%) In a statistical analysis of the data collected from the central site since 1994 (Longmore 2005), all fluxes except dissolved oxygen and carbon dioxide (Fig 24) varied significantly with sample year. As with the

SEES Head Technical Report: Nutrient Cycling 42

Hobsons Bay site, differences were not part of a consistent trend over time. Denitrification efficiency was lower in autumn 2004, 2005 and 2006 than in the previous spring, with the differences particularly marked for the central bay site in 2005-2006 (Fig 25). The data indicates a seasonal depression, but full recovery by the next spring. As a consequence, fluxes in 2005 were not different statistically (p>0.05) from those in 1995, 1996, 2003 or 2004, but they were significantly different to those in 1994.

Carbon dioxide flux, Central

25 ) -1 d -2 20

15 flux (m mol Cm mol (m flux

2 10 TCO 5

0 Sp Su Wi Su Sp Au Sp Au Sp Au Sp Au 1994 1995 1995 1996 2002 2003 2003 2004 2004 2005 2005 2006

Figure 24 Variation in carbon dioxide flux (mean ±±± 1 SD) in central Port Phillip Bay, 1994-2006 (Nicholson et al. 1996; Berelson et al . 1998; Longmore 2005).

Denitrification Efficiency, central

120

100

40 Denitrification Efficiency (%) EfficiencyDenitrification 20

0 Sp Su Wi Su Sp Au Sp Au Sp Au Sp Au 1994 1995 1995 1996 2002 2003 2003 2004 2004 2005 2005 2006

Figure 25 Variation in denitrification efficiency (mean ±±± 1 SD) in central Port Phillip Bay, 1994- 2006 (Nicholson et al. 1996; Berelson et al . 1998; Longmore 2005).

SEES Head Technical report: Nutrient Cycling 43

Two mechanisms by which denitrification efficiency could fall are: a failure in nitrification, or a failure of denitrification. Nitrification could fail if oxygen concentrations in the sediments became limiting (too low), while denitrification could fail if oxygen concentrations in the sediment were too high, or if the mechanism(s) by which nitrate is transported from oxygen-rich to oxygen-poor zones in the sediment failed. Higher nitrate fluxes accompanied the decline in denitrification efficiency in central PPB in autumn 2004, 2005 and 2006 compared to the previous spring, indicating the partial failure of denitrification, rather than of nitrification. Even so, oxygen and carbon dioxide fluxes were not particularly high. This suggests that a decline in the transport of water through the sediment (e.g. from a change in fauna living in or on the sediment) is a more likely explanation for decline in denitrification efficiency than is a change in supply of carbon to the sediment. The weakness in this argument is that the declines in denitrification appeared to be seasonal (in autumn), and there was no evidence of an impact in spring. In spring 2003, 2004 and 2005, denitrification efficiency at the central site was higher than the historic median, indicating full recovery between autumn and the next spring. Examination of the chamber Cs spike dilution data does not support the argument of a decline in bio-irrigation (Longmore, unpublished); in fact, the evidence suggests that bio-irrigation was higher in autumn 2005 and 2006 than in the previous springs. Such variations could also arise from increasing heterogeneity in the sediment, so that they are not seasonal variations, but rather random “noise”. Recent studies (Dr Jeff Ross, unpublished) have indicated that the impact of Sabella on denitrification depends on how the Sabella are distributed: clumps have a different impact to the same number of animals more evenly distributed. Presumably the same may apply to the many infauna in the Bay whose role in denitrification remains unknown. The power analysis on which benthic flux sampling effort is based (Longmore and Gason 2001) assumes that the distribution of fauna which play a role in denitrification does not change over time. Increasing variability in fauna (if it exists) means that the power to detect a specific level of change at current sampling effort will be lower than expected. Denitrification efficiency was high (means 81-97%) at the other sites in northern Port Phillip Bay sampled for the PPBES, with lower efficiency in shallow waters near the Western Treatment Plant, off the Patterson River and just outside Hobsons Bay (Nicholson et al . 1996). While Ross et al . (unpublished) found similar efficiencies in shallow waters (~ 12 m) along the eastern coast near Beaumaris and Seaford, they found lower efficiencies (means 47-63%) in 17 m deep waters off the same sites. Because of the variability in fluxes between replicate deployments at the same site, the apparent differences with depth were not statistically significant. However, if the differences are real, they do not match the previous measurements, which indicate lower efficiencies in shallower waters close to the known nutrient inputs. Ross et al . (unpublished) also found a very low efficiency (mean 33%) at an 8 m deep site east of Altona Bay. Denitrification efficiency at a site just east of the Port Melbourne Channel (Longmore 2006b) and on the southern section of the Port Melbourne DMG (Longmore unpublished) was generally high (a mean of 96% near the channel and 85% in the southern part). A lower efficiency (mean 68%, range 46-82%) in the northern part of the PoM DMG in March 2005 could be attributed to dredged material placement some months before the measurements, but efficiency had recovered (mean 84%, range 75-91%) 11 months later in Feb 2006 (Longmore unpublished).

SEES Head Technical Report: Nutrient Cycling 44

Summary for north of the Bay.

Key nutrient cycling processes:

• Benthic nutrient fluxes in central Bay are much lower, and denitrification efficiency much higher (at ~ 85%), than in Hobsons Bay; • There are significant differences in denitrification efficiency between sites in the area, with lower efficiency often observed closer to shore; • Because of the high efficiency, and large area, this area contributes significantly to the capacity of the Bay to denitrify.

Key factors affecting processes:

• Strong nutrient gradients exist in the water column in this area, especially near the north-east coast, Werribee and Hobsons Bay; • The Yarra plume generally travels down the east coast; • Nutrient concentrations and plankton biomass are low in the centre; • Spatial differences exist in the timing of peak phytoplankton biomass- highest in winter off the Werribee coast and in the centre, and highest in summer off Hobsons Bay; • A response was observed in the centre from the February 2005 flood; • There are significant differences in fluxes between years at the one site, and apparently between spring and the next autumn, the reasons for which are not yet clear; • Lower (at ~70%) denitrification efficiency found in the northern part of the PoMDMG may relate to recent disposal on the site, but if so, had recovered within a year; • Port Melbourne Channel sediments are nutrient-poor, compared to those from Hobsons Bay ; • The immediately released ammonium load in the sediments to be dredged is only about 10% of previous estimates, while that which becomes available over time is similar to previous estimates.

SEES Head Technical report: Nutrient Cycling 45

5.3 South of the Bay

The area under consideration here includes all waters on and south of the Sands, excluding the Entrance (Fig 26). Water depths vary from > 20 m in the north, to intertidal in some parts of the Sands. It includes all of the South Channel, and the proposed South East DMG.

Figure 26 Southern project area. The southern waters, including the South Channel, are low in nutrients and low in plankton biomass, but bioavailable nutrients from the dredge hopper are probably retained in the system long enough to be taken up by phytoplankton, rather than flushed to Bass Strait. South Channel sediments contain low concentrations of elutriated nutrients, with concentrations increasing toward the bottom (4-5 m deep). The bioavailable nutrient load from the South Channel is small in relation to other inputs, despite the large volume of sediments to be dredged. Little bioavailable N is expected to be released from particles after suspension in the dredge plume. Denitrification efficiency varies widely with location in the area, and over time at one site in the Sands, but is lower in shallow near-shore areas than in deeper waters. However, the high efficiency measured at some sites indicates this area also contributes significantly to the denitrification capacity of the Bay. Recovery of denitrification on a DMG appeared complete within 18 months of dredged material placement. Water column processes

Water movement is dominated by the wind north of the Sands (Fig 5), and by the tides on and south of the Sands. Simulated particle tracks in the South and West Channels (Prytz and Heron 1999) indicated particle movement with the tide, but the particles did not move far from the point of release after a complete tidal cycle. The average tidal run in the channels was 5 km, and since the channels are longer than that, Prytz and Heron (1999) concluded that tidal movement contributes little to flushing of Port

SEES Head Technical Report: Nutrient Cycling 46

Phillip Bay. These simulations also suggest that nutrients released in the south may persist long enough (several tidal cycles) to be taken up by plankton, rather than dispersed or lost to Bass Strait. Longmore et al . (1996) monitored a site off Blairgowrie on a fortnightly basis from 1990-97. EPA also monitored a site near Dromana on a nominal monthly basis from 1989 to 1995. These data indicate the same pattern of differentiation between N species as measured further north, that is, dissolved organic N concentrations greatest, followed by particulate N, ammonium and oxidised N (Fig 27). Oxidised N concentrations were often close to detection. Chlorophyll a concentrations from the same site were generally less than 1.5 µg L -1 (Fig 28), with the largest spike in November 1992 coinciding with a spike in particulate N (most likely due to N held in the plankton), but not with ammonium or oxidised N.

Blairgowrie

10 M) µ µ µ µ 8 Ammonium Oxidised N Organic N 6 Particulate N

Concentration( 4

0 Jan-90 Jan-91 Jan-92 Jan-93 Jan-94 Jan-95 Jan-96 Jan-97 Jan-98 Date

Figure 27 Nitrogen species concentration in southern Port Phillip Bay, 1990-97 (Longmore et al . 1997).

Significant numbers of the bitter-taste diatom Rhizosolenia cf chunii have been observed in the south, both at Blairgowrie (Nicholson et al . 1998) and more recently at Dromana ( Appendix 1). Other potentially toxic algae ( Pseudo-nitzscia spp, Dinophysis acuminata and Gymnodinium sp .) have also been observed at Dromana at levels exceeding the VSQAP guidelines.

SEES Head Technical report: Nutrient Cycling 47

Blairgowrie

4.5

3.5 ) -1 3 g L g µ µ µ µ (

a 2.5

1.5 Chlorophyll Chlorophyll 1

0.5

0 Jan-90 Jan-91 Jan-92 Jan-93 Jan-94 Jan-95 Jan-96 Jan-97 Jan-98 Date

Figure 28 Chlorophyll a concentrations from Blairgowrie, 1990-97 (Longmore et al . 1997).

Benthic processes

No cores were collected from the southern area of Port Phillip Bay for the PPBES. Kowarsky (2002) sampled sediments in South Channel (15 samples), with a geometric mean total organic carbon concentration of 1.8%. Mean total organic carbon concentration was lower at depth (1.2% at 0.8 m depth) in cores from South Channel, but the magnitude and sign of the differences between surface and bottom sediments differed considerably for each core. Three cores from the South Channel were analysed for the CDP EES (SKM 2004). Two gave contrasting changes in ammonium concentration with depth. One (borehole 51) showed substantial ammonium concentrations at 2.9-3.0 m depth, while the other (borehole 52) had undetectable ammonium concentrations at depth. These data suggested the sediments in the South Channel are variable in ammonium concentration, and that we cannot assume that deep (> 1 m) sediments are nutrient-poor. Sediments were collected from the hopper of a dredge during the Trial Dredge Program (TDP), while dredging in South Channel. Ammonium, oxidised N and dissolved organic nitrogen (DON) concentrations were measured in the aqueous phase. Wet sediment was suspended in a column of seawater, and particles of sizes 2-63 µm sampled. These were analysed for total nitrogen concentration, to determine the distribution of N across the particle size range. Sediments from the South Channel were dominated by coarse fractions: more than 95% of the sample weight was in particles > 63 µm in diameter (Longmore 2006a). These are expected to settle very close to the area of discharge. Concentrations of ammonium and DON in the aqueous phase from the South Channel were much higher than the limited sampling during the PPBES led us to expect for sandy sediments. Nitrate formed only a minor part of the total dissolved N concentration. The dissolved N load estimated to have arisen from the TDP of 3.7 ± 0.6 tonnes N in the south, was 148% of that predicted in the EES (scaled to the volume dredged in the TDP). With the larger than expected DON load in the south (relative to ammonium), the issue of DON bio-availability was addressed by additional experiments . Nine cores from the South Channel, and three from the SE DMG, were analysed for elutriates (Longmore 2006c). All of the elutriated nutrients were low in concentration compared to sediments from the other

SEES Head Technical Report: Nutrient Cycling 48

project areas. Mean concentrations were highest at 4-5 m, rather than near the surface. Bioavailable nutrient loads expected from the hopper and/or overflow from the planned dredging in the South Channel are shown in Table 11. They indicate lower ammonium and DON loads than estimated for the EES.

Table 11 Estimates of dissolved nutrient load (tonnes, mean ±±± 95% CI) in sediment to be dredged from the South Channel (SKM 2004; Longmore 2006a; Longmore 2006c).

Period Ammonium Oxidised N DON Phosphate Silicate 2003 23.0 46.0 2005- TDP 14.4 ± 4.2 0.52 ± 0.33 34.0 ± 18.9

2006 coring 9.6 ± 6.2 0.08 ± 0.00 11.3 ± 2.7 1.1 ± 0.7 29.8 ± 6.2 Mean DON concentration measured in the elutriates from the recent coring was only about 25%, and oxidised N only 10%, of the mean from the TDP sampling. This accounts for the difference between estimates in Table 11. Interestingly, the cores collected from the SEDMG were richer in elutriated nutrients than the cores from the South Channel, suggesting that the sediments dredged and placed at the DMG during the TDP were somewhat richer in nutrients than the sediment sampled recently. The total N loads calculated for the sediments removed in the TDP (151 ± 25 t N), when extrapolated to the volume expected to be dredged in the same areas for the capital dredging program, were 60% of those predicted for the EES. Conversely, the total N load calculated from the most recent sediment sampling is more than twice that (5,000 t; Longmore 2004) estimated for the EES (Table 12). These differences indicate the bias that can be introduced to estimates by using small sample numbers.

Table 12 Total nutrient load to be mobilised by dredging in the South Channel (Longmore 2006c).

Area Load (t, mean ±±± 95% CI) In situ volume 15.04 dredged (million m 3) Mean N load 7,975 ± 5,100 Mean P load 4,350 ± 1,800 Mean Si load 17,200 ± 9,100

Contrary to expectations, sediments in the South Channel from the TDP were richer in total N per unit of dredged sediment than those from the Port Melbourne Channel (Longmore 2006a). This was attributed to the Port Melbourne Channel sediments being organically-poor clays. Those particles carrying the greatest proportion of the N load were in the range 2-16 µm; such particles may remain suspended for up to 100 hours before they reach the seafloor. The potential therefore exists for some fraction of the particle-bound N to become bio-available, and this issue has now been addressed. Many of the samples from the South Channel showed ammonium uptake over time, compared to release for the other areas. The only samples from the South Channel showing a net release of ammonium over 10 days were from below 2 m at the SE DMG (and assumed to be recently placed dredge material). Changes over 10 days, extrapolated to the dredge volume for the South Channel, are listed in Table 13.

SEES Head Technical report: Nutrient Cycling 49

Table 13 Estimates of bioavailable nutrient load (tonnes, mean ±±± 95% CI).after 10 days in sediment to be dredged from the South Channel (SKM 2004; Longmore 2006c)

Period Ammonium Oxidised N DON Phosphate Silicate 2003 7.0* 3.0*

2006 coring -2.6 ± 24.9 4.61 ± 15.5 8.5 ± 10.4 0.2 ± 2.2 22.6 ± 13.8

In summary, the sediment chemistry indicates that South Channel sediments may release relatively small quantities of bioavailable N in overflow at the dredge site, and very little from particles in the dredge plume. This is despite a large total N pool, which supports the view that the vast majority of the total N mobilised by dredging in the South Channel will not become bioavailable., and can be considered inert from a nutrient cycling point of view.

Direct measurements of benthic nutrient fluxes in southern Port Phillip Bay have been made at six sites, one for the PPBES, four for the Channel Deepening Project, and one for the ARC Linkage program. The PPBES site was sandy, on the northern edge of the Sands (Berelson et al . 1998). Fluxes were much lower there than at all other sites measured during the PPBES. However, denitrification efficiency was high (90 ± 5 %). Measurements in the sandy Capel Sound DMG (Longmore et al . 2004) and from a muddy site in the proposed South-East DMG (Longmore unpublished) also indicated high efficiency (90-95%). Dredge material had been placed in the Capel Sound DMG about 18 months before the flux measurements were taken, indicating that denitrification efficiency, if impacted by disposal, had recovered to extremely high levels within 18 months. Lower efficiencies have been observed in shallow waters off Mornington (Ross et al . unpublished; 55 ± 45 %) and off St Leonards (Longmore unpublished; 75 ± 7 %). Detailed measurements were made at a site west of Hovell Pile on the northern edge of the Sands, for a MPB shading experiment in April 2005, and again as part of the Trial Dredge Program in September 2005 (Longmore 2006b). Benthic fluxes were below the median of previous measurements at the Hobsons Bay and central sites, and again presumably reflect a relatively low supply of organic matter to the sediments. There was no evidence of a substantially lowered ambient DO concentration at the sediment surface. Silicate fluxes correlated strongly with carbon flux, suggesting diatoms (either planktonic or MPB) were contributing to the organic matter being recycled. Denitrification efficiency in September 2005 was similar to that at the other sites in the south (94 ± 1 %). Conversely, the site showed quite different behaviour in April 2005, with high ammonium fluxes and low denitrification efficiency (mean 14%), at a relatively low respiration rate. The reasons for this difference are not yet clear. Denitrification efficiency was zero for two of the replicates, because the N 2 flux was negative (apparently indicating N 2 uptake by the benthos). Edmunds et al . (2004) found low numbers of cyanobacteria at five sites in the south, but only one in the north, and so N fixation is a theoretical possibility in the south. For the other two replicates, the efficiency was very low (19 and 37 %). This site is in an area that is both biologically active and subject to large-scale sediment transport, and it is possible that one or other of the deployments was made soon after the site had been disturbed. Longmore (unpublished) found an efficiency of 90% at a site in the SE DMG, while Ross et al . (unpublished) have observed much lower efficiency (55%) in shallower inshore waters nearby. Longmore (unpublished) has also found lower efficiency (70%) at a shallow site off St Leonards.

SEES Head Technical Report: Nutrient Cycling 50

Summary in the south

Key nutrient cycling processes

• Denitrification efficiency varies widely with location in the area, and over time at one site in the Sands; • Denitrification efficiency is lower in shallow near-shore areas than in deeper waters; • Despite the PPBES model predicting low denitrification efficiency on and south of the Sands, because of a predicted low carbon supply, the high efficiency measured at some sites indicates this area also contributes significantly to the denitrification capacity of the Bay.

Key factors affecting processes

• Ineffective flushing leads to nutrients entering the area retained in the system long enough to be taken up by phytoplankton, rather than flushed to Bass Strait; • The southern waters, including the South Channel, are low in nutrients and low in plankton biomass; • Recovery of denitrification on a DMG appeared complete within 18 months of dredged material placement. • South Channel sediments contain low concentrations of elutriated nutrients, with concentrations increasing toward the bottom (4-5 m deep); • Sediments recently placed in the SE DMG contained higher elutriate nutrient concentrations than those from the area dredged, indicating rapid replacement of pore water nutrients lost during dredging; • The bioavailable nutrient load from the South Channel is small in relation to other inputs, despite the large volume of sediments to be dredged and the large total N load.

SEES Head Technical report: Nutrient Cycling 51

5.4 The Entrance

Under consideration here is the area between Point Lonsdale, Queenscliff and Point Nepean (Fig 29). It includes rocky bottom, and the canyons up to 90 m deep that partly lie beneath the shipping channel. Low ambient nutrients, low phytoplankton biomass, high turbulence, high exchange and lack of muddy or sandy substrate suggest that nutrient cycling at the sediment surface will be insignificant in the area; most detrital material (eg. fragmented macroalgae) is likely to be transported out of the area.

Figure 29 The Entrance project area.

SEES Head Technical Report: Nutrient Cycling 52

Water column processes

Water circulation is dominated by the tides, and to a lesser extent by the other factors (eg. winds, air pressure) that affect water level in Bass Strait. There is little data available on water quality at the Entrance to Port Phillip Bay. Longmore et al. (1996) collected measurements from a site just inside the Entrance 21 times over a year. The relative concentrations of each of the N species was similar to that measured at Blairgowrie (Fig 30), though the average oxidised N concentration was higher at the Entrance than at Blairgowrie. This could indicate Bass Strait as an occasional source of oxidised N to the Bay. As with all other areas (except the Werribee coast), dissolved organic N was the dominant form. The mean chlorophyll a concentration was ~ 0.6 µg L -1, the lowest mean of the four areas. Flushing is thought to be more effective in the area than elsewhere in the Bay, which may decrease the possibility of algal blooms, even if nutrient concentrations were high enough.

Entrance

7 M) µ µ µ µ 6 Ammonium Oxidised N 5 Organic N 4 Particulate N

Concentration ( 3

0 Nov-93 Mar-94 Jun-94 Sep-94 Jan-95 Apr-95 Jul-95 Date

Figure 30 Nitrogen species concentrations at the Entrance, 1994-95 (Longmore et al . 1996).

Benthic processes

No sediment samples have been collected from the Entrance for nutrient analysis.

SEES Head Technical report: Nutrient Cycling 53

No measurements have been made of nutrient cycling in the Entrance region. However, because the area is dominated by rocky surfaces, rather than muddy or even sandy sediment, there is little likelihood that the seabed plays any role in nutrient cycling.

Summary for the Entrance

• Low ambient nutrients, low phytoplankton biomass, high turbulence, high exchange and lack of muddy or sandy substrate suggest that nutrient cycling will be insignificant in the area; most detrital material (eg. fragmented macroalgae) is likely to be transported out of the area.

SEES Head Technical Report: Nutrient Cycling 54

5.5 Summary

The conceptual models (Figs 2, 3) and detail in this chapter demonstrate that the same key factors governing nutrient cycling in Port Phillip Bay apply in each Project Area. These factors include: • the supply of nutrients (with nitrogen as the key limiting nutrient); • rapid incorporation of nutrients into phytoplankton, followed by intense recycling in the water column, via zooplankton grazing and excretion, cell leakages etc; • deposition of nutrients in the sediment ( via dead cells and faecal pellets); • recycling of nutrients by microbes in the sediment; • oxygen mediation of the products of nutrient recycling; • the potential role of MPB in assimilating nutrients and altering the oxygen profile in the sediments; and • the potential role of infauna. However, the relative importance of each factor varies with Project Area and season. Key nutrient cycling processes

Phytoplankton production is seasonal, with a peak in summer. Baywide, about 40% of phytoplankton production falls to the sediment, where it drives benthic fluxes. Denitrification efficiency is inversely related to the rate of supply of organic matter. Carbon supply is lower in the deep central basin than in Hobsons Bay, and denitrification efficiency is higher (~ 85%). Carbon supply is also low in the southern region, and denitrification efficiency high. Denitrification efficiency is lower on the Werribee coast (but still higher than in Hobsons Bay).

Key factors affecting processes

The major nutrient inputs to the Bay are in the two northern Project Areas, from the Western Treatment Plant (WTP), on the west coast, and the Yarra River. The inputs vary seasonally, with highest WTP input in the winter, and highest Yarra input in spring-summer. The plumes from each of the major inputs spread according to the wind direction, with a high proportion of the nutrients assimilated by phytoplankton close to the sources. In the Southern Project Area, circulation is also wind-driven, with a large anti-clockwise gyre forming north of the Sands in westerly winds. Because of spatial variability in phytoplankton production, this supply of organic matter to the sediment also varies spatially. The benthic fluxes likewise vary seasonally, and they also vary spatially. Carbon fluxes are highest in summer in Hobsons Bay, and denitrification efficiency is lowest (~ 60%). Imposed on this spatial pattern are large, possibly seasonal (or even shorter), changes in denitrification efficiency in the highly mobile Sands area, and some other near-shore areas. Nutrient recycling in the water column has been ignored so far, because it results in the return to the water column of nutrients that are rapidly re-incorporated into phytoplankton. More importance has been placed on those processes that result in the permanent removal of nutrients from the system. However, we cannot ignore the critical pathway necessary to deliver nutrients from the catchment to the sediment prior to denitrification, that is, uptake by phytoplankton. Part of the reason the Bay is so efficient at denitrification is that a high proportion of the plankton fall to the sediment. They do this because they are relatively large diatoms. Shifts in phytoplankton community composition toward smaller cells, unless compensated for by increased grazing and increased faecal pellet production, may lead to a greater proportion of phytoplankton production being recycled in the water column. This would increase the supply of nutrients available for plankton growth, and a greater frequency of plankton blooms may ensue.

SEES Head Technical report: Nutrient Cycling 55

While ambient nutrient, chlorophyll and dissolved oxygen concentrations shed light on conditions important for nutrient cycling, with our current level of understanding and a generally well-mixed water column, it is not possible to use existing water quality monitoring information to infer denitrification efficiency. Concentrations in the water column do not necessarily reflect the rate of transformations of nutrients. For example, in Port Phillip Bay there is little evidence in the water column of dissolved oxygen depletion, even when it is zero just below the sediment surface. This is because the water column is usually well mixed. Burke (1999) demonstrated that measurements would have to be made within 100- 400 µm of the sediment surface to detect the benthic oxygen gradient. This is beyond the capability of most commercial field equipment. Similarly, while there is a link between primary production and benthic fluxes (Nicholson and Longmore 1999), spatial and temporal variability in chlorophyll measurements make it difficult to see such links in the field. A measure of primary production at much more meaningful spatial and temporal scales is required to advance our knowledge in this area, beyond that provided by models. Similarly, studies of sediment nutrients can shed light on some of the driving forces, but laboratory measurements under-estimate in situ fluxes (Macreadie et al . 2006). With our current level of understanding, the only reliable method of detecting changes in denitrification is by using benthic chambers.

Seasonality

Seasonal aspects of nutrient cycling have been noted above, and are summarised below (Table 14). Overall, from the point of view of potential impacts on algal blooms and denitrification, summer-autumn is the most sensitive period in the north, and winter is the most sensitive season in the south.

Table 14 Summary of seasonal issues with nutrient cycling

Area Seasonal issues Yarra River and Nutrients and chlorophyll concentrations often lowest in winter, and highest in Hobsons Bay summer. Floods more likely in spring-summer. Potential for algal blooms in general, and toxic algal blooms in particular, greatest in summer-autumn, and least in winter. Denitrification thought to be less efficient in summer than winter; benthic nutrient fluxes are highest in summer. MPB biomass is probably highest in summer. North of the Bay High N concentrations brought in by summer storms extend beyond Hobsons Bay down the east coast. Highest concentrations off the WTP occur in winter. Potential for algal blooms in general, and toxic algal blooms in particular, greatest in summer- autumn, and least in winter. MPB biomass is probably highest in summer. Impacts of floods not thought to extend to the centre, but denitrification efficiency is lower in summer-autumn throughout the area. South of the Bay Increased light and temperature in spring-summer leads to increased phytoplankton and MPB growth. The potential for algal blooms is greatest in summer, except for Rhizosolenia , which is most likely in winter. Denitrification efficiency is expected to be lower in summer than winter. The Entrance No issues identified

SEES Head Technical Report: Nutrient Cycling 56

6 Interactions Arising from the Project

We have identified ten mechanisms by which dredging may plausibly affect nutrient cycling and denitrification (Table 15). These will be discussed in turn.

Table 15 Plausible modes of dredging impact on nutrient cycling.

Initiating event Type of impact Effect of impact Removal of seabed Channels Loss of denitrification

Nutrient release from plume Increased phytoplankton Increased carbon supply to production sediment; Reduced nitrification; Decline in denitrification efficiency Nutrient release from plume Increased phytoplankton Stimulation of toxic algal blooms. production Turbidity from plume Reducing phytoplankton Reduced carbon supply to production sediment; Enhanced nitrification, but suppression of denitrification. Turbidity from plume Reduced MPB production Increased nutrient flux from sediment; Reduced denitrification efficiency Sediment from plume Burial of MPB Increased nutrient flux from sediment; Reduced denitrification efficiency Sediment from plume Burial of infauna Reduced nitrification; Reduced denitrification efficiency Toxicants from plume Reduced grazers Increased carbon supply to sediment; Decline in denitrification efficiency Toxicants from plume Toxic effect on infauna Reduced nitrification; Reduced denitrification efficiency Exotic species introduced Reduce bio-irrigating fauna Reduced nitrification; Reduced denitrification efficiency

SEES Head Technical report: Nutrient Cycling 57

The first plausible impact is the loss of effective denitrification surface by removal of seabed (Fig 31). This impact is plausible because dredging necessarily removes seabed, and we know that denitrification is carried out in the sediment. Removal of seabed will remove the bacteria responsible for nitrification and denitrification, and also the infauna that irrigate the sediment. Nitrogen in the sediment that would normally be lost to the atmosphere by the sequence organic N ammonium nitrate N 2 gas is instead returned to the water column as ammonium, where it may promote further algal growth. If algal growth is excessive, bottom-water anoxia may ensue, preventing nitrification by a re-established bacterial population, and leading to even more ammonium released. This impact applies in the dredged areas (the channels). How important the impact is depends on how much the channel bed contributes to the denitrification capacity of the Bay, and how quickly recovery will occur once dredging ceases.

Figure 31 Removal of seabed by dredge prevents denitrification. Symbols courtesy of the Integration and Application Network ( ian.umces.edu/symbols/ ), University of Maryland Center for Environmental Science.

SEES Head Technical Report: Nutrient Cycling 58

The next plausible impact is that of nutrients released from the dredge plume, leading to increased phytoplankton production (Fig 32). Since phytoplankton growth is nutrient-limited in most of the Bay for at least a part of the year, we could expect dredging to lead to increased plankton production. The next link in this chain is an increase in plankton biomass settling to the seafloor, leading to a reduction in oxic condition of the sediment, a consequent reduction in nitrification, a decline in denitrification, and increased release of ammonium. This can then stimulate further plankton growth. The nutrients released could be bioavailable dissolved nutrients in the aqueous phase in the dredge hopper, or discharged in overflow. They could also be those released over a longer period of time from suspended material in the water column, or once the particles settle to the seabed. Therefore, the area of nutrient release could be at the dredge site, anywhere under the plume, anywhere fine particles from the plume may be transported over several days, or at the DMG.

Figure 32 Nutrients from the dredge plume increase phytoplankton production. The area of plankton growth will be wherever the dissolved nutrients disperse to within a few days. Phytoplankton can take up nutrients in excess of their immediate needs, so that growth may be observed some distance from the area of uptake, or some time after the plume clears. The area of plankton deposition will be close to the area of growth, if diatoms dominate, or further dispersed if smaller cells dominate. The key to local deposition of the plankton is the type of plankton stimulated- diatoms or other. Hence, this pathway is complicated by the possibility that nutrient stimulation may lead to altered phytoplankton community composition, and thereby altered balance between nutrients recycled in the water column, and from the sediment.

SEES Head Technical report: Nutrient Cycling 59

An extension of this potential impact is the situation where the plankton that blooms is toxic (Fig 33). This second impact is plausible because toxic blooms have already been observed in Port Phillip Bay, and have followed storm flows in the Yarra. Though there is a common perception that harmful algal blooms are occurring more frequently around the globe now than in past decades, a direct link between harmful algal blooms and eutrophication has been elusive (Parsons and Dortch 2002). In part, this has been due to lack of regular phytoplankton monitoring in coastal areas prior to blooms being detected. Even though toxic blooms have occurred in northern Port Phillip Bay, and the toxins found in mussels, the next steps, which include transfer of the toxins to higher levels in the food web (fish, birds, mammals, man) have not been proven in the Bay.

Figure 33 Nutrients from the dredge plume cause toxic algal blooms.

The action of the dredge plume leading to the suspension of toxic algal cysts in the water column can only enhance the plausibility of this pathway. Impact could therefore involve stimulation of a bloom, or support of an existing bloom.

SEES Head Technical Report: Nutrient Cycling 60

The next plausible impact is the opposite of the first: rather than nutrient stimulation of plankton growth, in this case turbidity from the dredge plume leads to a loss of photosynthetic capacity, so that phytoplankton biomass declines (Fig 34). Such a process has been observed in naturally turbid estuaries, in the area of maximum turbidity (Irigoien and Castel 1997). This may impact higher trophic levels, though that issue is not pursued here. To lead to an impact on denitrification, the carbon load to the sediment must decline so much that the sediment remains oxic throughout the zone involved in nutrient cycling. Denitrification is therefore prevented, and nitrogen returned to the water column as nitrate. This may stimulate plankton growth, but not until the light climate improves. The area of impact will be directly under the plume, although the impact may be observed at quite high dilution of the plume, particularly on occasions when plankton growth is already light-limited. Any such impact on the oxic status of the sediment may be partly offset by oxygen demand from the dispersed sediment.

Figure 34 Turbidity from the dredge plume reduces phytoplankton production.

SEES Head Technical report: Nutrient Cycling 61

A related impact is that of turbidity from the plume reducing MPB production (Fig 35). In this case, the impact on production is on the seafloor, rather than in the water column. If microphytobenthic productivity declined, we might expect to see nutrients that would otherwise have been involved in growth now released to the water column. Similarly, if MPB plays a role in denitrification, this could decline. If the fluctuations in oxic state (oxygen produced by MPB during the day, or consumed at night) are significant in supporting denitrification, the loss of production could also lead to a decline in denitrification. The impact would last as long as the plume persists, perhaps with some recovery time as well.

Figure 35 Turbidity from the dredge plume reduces microphytobenthos production.

SEES Head Technical Report: Nutrient Cycling 62

The next impact is almost identical to the last, but in this case MPB production is lost because the cells are buried by sediment, rather than shaded by turbidity in the water column (Fig 36). For all of the reasons advanced above, denitrification may be lost. This impact would be restricted to a small proportion of the plume (principally along the channel edges), and the DMG, and the period of impact would presumably be longer than that induced by turbidity. Assuming burial is so deep that all viable cells are killed, recovery could only occur once new cells arrive, either by germination (for those growing from cysts) or advection from the water column.

Figure 36 Sediment from the dredge plume buries microphytobenthos.

SEES Head Technical report: Nutrient Cycling 63

The next plausible impact involves the burial of infauna by dredged material (Fig 37). Parslow and Murray (1999) concluded that the comparisons of Nicholson et al. (1996) between advective and diffusive fluxes indicated about half of the organic matter was decomposed in the surface few millimetres, while the rest was decomposed up to 50 cm deep. The principal mechanism for the deep decomposition was the action of bio-irrigating and bioturbating infauna, which introduce both oxygen and organic matter deep within the sediment. Even noting the relatively weak correlation between denitrification and bio- irrigation observed by Berelson et al . (1998), we have accepted that infauna probably play an important role in the high efficiency with which the Bay sediments denitrify. If we likewise accept that about half of the denitrification is attributed to deep decomposition of organic matter, it follows that burial of the infauna responsible for the bioirrigation/bioturbation will lead to a partial loss in denitrification. This would apply in the same area as the last example: in a restricted part of the plume, and on the DMG. It would persist for as long as it takes for infauna to re-establish.

Figure 37 Sediment from the dredge plume buries infauna.

SEES Head Technical Report: Nutrient Cycling 64

The next plausible impact is one in which toxicants mobilised by the dredge affect zooplankton that would otherwise be grazing on phytoplankton (Fig 38). We have noted in Chapter 5 that about 60% of nitrogen cycling occurs in the water column, principally because of the grazing of zooplankton. Loss of zooplankton will lead to a loss in the rain of faecal pellets to the bottom, but more importantly (here) will also lead to an increase in plankton biomass. Once that occurs, the impact is the same as noted above for stimulation by nutrients in the plume: increased supply of organic matter to the sediment, decreased nitrification and /or denitrification, and an increase in nutrient supply to the water column. This then stimulates further algal growth.

Figure 38 Toxicants from the dredge plume reduce zooplankton grazing on phytoplankton.

SEES Head Technical report: Nutrient Cycling 65

The next impact, that of toxicants on infauna (Fig 39) is similar to the previous plausible impact (Fig 38), in that it leads to a reduction in denitrification. However, in this case it is due to the loss in bio- irrigation/bioturbation (similar to burial of infauna by the dredge material- Fig 36), rather than increased supply of organic matter to the sediment. The impact is presumably restricted to the area of heaviest sedimentation.

Figure 39 Toxicants from the dredge plume affect infauna.

SEES Head Technical Report: Nutrient Cycling 66

The last plausible impact is that of invasion by exotic infauna, either new to the Bay as a whole, or translocated from one area of the Bay to another (Fig 40). The exotic fauna impact by occupying habitat formerly occupied by the infauna we think are important to denitrification, or by consuming them. The vector of distribution is the hull, bilge, hopper and other spaces likely to hold live fauna, eggs or cysts. The loss of the fauna will lead to the partial loss of denitrification.

Figure 40 Invasion by exotic species alter infauna.

In addition to the plausible modes of impact listed above, there are also a number that are not considered feasible. These include the enhancement of MPB by nutrient release from the plume, and effect of toxicants on phytoplankton and MPB. The former is excluded from further consideration because MPB rely almost exclusively on nutrients in the sediment, where concentrations are usually much higher than in the water column (Lukatelich and McComb 1986; Nozais et al . 2001; Welker et al. 2002). The latter is excluded from further consideration because concentrations of toxicants known to affect phytoplankton (and by inference MPB) are much higher than those likely to be found during dredging (Nayar et al . 2004; Muller et al . 2005; Perez et al . 2006).

SEES Head Technical report: Nutrient Cycling 67

7 Assessment of Impacts

This section outlines the results of nutrient cycling studies carried out for the EES, TDP and SEES. Specifically, it provides: • A summary of impacts predicted for the EES; • The results of SEES studies to better define the nutrient load likely to be generated by dredging; • The results of an SEES study to measure impact of trial dredging on MPB and denitrification; • An estimate of the spatial impact of nutrient mobilisation, based on the revised turbidity plume modelling; • Comparison of dredging load with existing loads; • A description of the risk assessment process followed; • The risk assessment. The latter six points directly address requirements in the SEES Assessment Guidelines.

7.0 New evidence of impact

There are a number of ways in which we can assess the impact of dredging on nutrient cycling. We can assess the likely impact from channel deepening on nutrient cycling by observing the output of the PPBES model under scenarios involving increased inputs. The model (Murray and Parslow 1999) predicts a Bay-wide linear response to an additional 500 t N from the WTP and the Yarra River. All fluxes increased by the same proportion (15% off Werribee, and 30% off Hobsons Bay). We could therefore assume that, providing the dredging equivalent nutrient input is in the order of 500-1,000t N, the response of the Bay will be linear. However, it is likely that the impact may be different at local scales to the model predictions on a Bay-wide scale. When modelled on a regional basis, Murray and Parslow (1997) concluded that increased nutrient loads from the Yarra will lead to prolonged blooms in Hobsons Bay, and more intense blooms along the eastern shore. The model responses are larger than linear, because the denitrification efficiency, which is already low in Hobsons Bay, is reduced further. The risk of prolonged oxygen depletion in bottom waters increases with load. Once impacted, we also need to know how the system may recover. Kemp et al. (2005) state that recovery of nutrient cycling from anoxia differs between N and P. Phosphorus cycling (because it is almost entirely controlled by chemical processes) can be restored within hours of the alleviation of anoxia, but nitrifying bacteria take weeks to months to recover from an anoxic episode (Kemp et al. 2005). There may also be a longer-term impact on cycling of N and P from the effects of anoxia, physical removal of invasions of exotics on infauna. In addition to model predictions, we have the results of studies carried out for the EES (2003-04), before and during the Trial Dredge Program (2005) and more recently (2006).

7.01 ESS studies and conclusions

In order to determine the potential for the dredging/disposal process to remobilise nutrients, and change nutrient cycling and denitrification processes, the process of dredge material placement was simulated in

SEES Head Technical Report: Nutrient Cycling 68

the laboratory. This assessment, together with field denitrification studies, led to the following conclusions: • Placement of dredge material in the Port of Melbourne DMG is likely to lead to an initial nutrient release, especially of ammonium, and oxygen consumption, but no large change in nutrient fluxes from sediments. Sulphide found in some Yarra River and PoM Channel sediments may exert a further oxygen demand. In contrast, placement of dredge material in the southern DMG is likely to lead to only minor release of nutrients initially, but a much more significant change in denitrification efficiency • from sediment analysis, the total ammonia-nitrogen load associated with the sediments to be dredged was estimated as 157 tonnes from the northern areas, and 23 tonnes from the southern areas. On a Bay-wide basis over the expected period of dredging, these were very small loads compared to the external inputs of approx. 5-15 tonnes d -1 bioavailable N from the WTP and Yarra River (Harris et al . 1996). These N loads equated to an input on the dredge material grounds of 9 and 2 g N m -2 in northern and southern DMGs respectively, compared to the current bay-wide external load of approximately 3.5 g N m -2 . • assuming dredging of Yarra River sediments occurred without overflow, in order to achieve background ammonium concentrations during disposal of Yarra River sediments, the aqueous phase in the dredge hopper needed to be diluted a minimum of 10 to 85 times on disposal at the PoM DMG, and to reach background silicate concentrations, dilution of 1 to 20 times would be required • the dredge hopper contents would need to be diluted up to five fold to fall below the ANZECC/ARMCANZ water quality guidelines for ammonia toxicity • benthic fluxes in the south-eastern DMG were expected to increase to a level which is equivalent to a doubling of external N loads . If the nutrients released by dredge material placement remain within the DMG, and are taken up by phytoplankton before they can be more widely dispersed, an approximate doubling of the plankton biomass (from ~ 1 to ~ 2 µg chlorophyll a L -1 ) can be expected. However, the plankton is also likely to be dispersed over a much wider area by normal wind-driven mixing, so that an increase in biomass may be difficult to detect • during the Yarra/Hobsons Bay dredge material placement, at least an 8-fold dilution would be necessary to avoid short-term loss of oxygen to below the SEPP limit (90 per cent saturation) in the PoM DMGs • considering the amount of light currently reaching the sediment surface at the PoM DMG and the proposed new south-eastern DMG, any impact of increased turbidity in these areas from dredge material placement would be marginal, as there was no evidence that the MPB were very productive in these areas • monitoring during and following the Geelong Channel Improvement Program (CIP) found that there was no live macrobenthic infauna in freshly placed material. Many species of infauna had recolonised the sediment within two months, and by six months there was no significant difference in population numbers or number of taxa between the DMG and a reference site. However, some changes in community persisted for the next four years, perhaps because sediment characteristics were changed at the DMG. This suggested that disposal to an existing DMG may have a lesser impact than disposal to a new site. The initial risk assessment concluded that dissolved oxygen depletion and ammonia toxicity are not likely to be major concerns, providing an effective dilution of at least 8:1 was achieved for the liquid phase on dredge material placement. In addition, normal water current movement was likely to spread the plankton which could grow as a result of any nutrient release over wide areas of the Bay, diluting the impact to the extent that it should not be detected.

7.02 SEES baseline study

Baseline measurements of nutrient fluxes and denitrification have been carried out in autumn 2005, spring 2005, late summer 2006 and winter-spring 2006. Results available so far have already been discussed in the appropriate sections in Chapter 5, and were within the ranges previously measured in Port Phillip Bay. In particular, the expected relationships between dissolved oxygen and ammonium

SEES Head Technical report: Nutrient Cycling 69

fluxes, and dissolved oxygen flux and denitrification efficiency, were observed at all sites except 2719 on the Sands, at which large changes in fluxes and denitrification have been observed. The reasons for this difference are not yet clear, but may be clarified by examination of subsequent baseline measurements.

7.03 SEES MPB experiments

While it has been hypothesised that up to 50% of the effective denitrification may be promoted by single celled microscopic plants (microphytobenthos, MPB) living on and in the top few mm of the sediment (Murray and Parslow 1997), no measurements of the role of MPB in denitrification exist for Port Phillip Bay. There is a concern that turbidity or burial from dredge plumes may smother or shade the MPB, negating its role in denitrification. An experimental study of the links between light attenuation, MPB and denitrification, at site 2719 on the northern edge of the Sands, was carried out in early 2005, using two densities of shadecloth to simulate a dredge plume. The experiment provided extremely interesting data. No significant change in respiration, biomass or photosynthetic characteristics of the MPB was detected, after one week of shading at 60% and 90% light attenuation. This indicates that MPB did not adapt to the lower light levels over a week, but equally that biomass did not decline. The photosynthetic parameters measured were all within the range previously measured at similar depth in Port Phillip Bay. Denitrification efficiency was low in both control and treatment areas, with no apparent effect from shading. This raised the possibility (admittedly from only one experiment at one site) that denitrification is naturally inefficient in the Sands, and therefore that turbid plumes from dredging in the south may have little impact on the capacity of the Bay to denitrify. This experiment was repeated, with replicated measurements before, during and after dredging at two sites in Port Phillip Bay, during the Trial Dredge Program in 2005 (Longmore 2006b). The sites were 6004, about 300 m east of the Port Melbourne Channel, and site 2719, west of the Hovell Pile. Analysis of variance indicated statistically significant differences between sampling periods at site 6004 for: • MPB biomass, • the light level required to saturate production (I k), • dissolved oxygen, carbon dioxide, nitrate, silicate and dissolved inorganic nitrogen (DIN) fluxes; and at site 2719 for: • ammonium, phosphate and DIN fluxes.

Only changes in DO flux, MPB biomass and nitrate flux at 6004, and in ammonium and DIN fluxes at 2719, appeared to be related to dredging. There was no statistically significant difference in denitrification efficiency with sampling period at either site. An analysis of the processes involved in nutrient cycling also indicated little impact from dredging at these sites.

No other MPB properties were affected at the northern site during dredging, but total MPB productivity declined because of the reduction in available light and loss of biomass. Several alternatives may explain the failure to observe a statistically significant change in denitrification efficiency: (i) MPB do not play a role in denitrification; (ii) the period of impact was too short to affect MPB properties (other than biomass and I k), or (iii) the degree of replication was insufficient to pick up small changes. The decline in MPB biomass in the north could simply be due to burial by dredge material, rather than due to loss of productivity. Sundback and Graneli (1988) found that microphytobenthos biomass changed little over several weeks of darkness, and increased soon after exposure to light. However, Edmunds et al. (2005) also noted a decline in MPB biomass in the north prior to and during dredging, which was part of a regional decline in biomass unrelated to dredging. They believed that dredging led to a delay in recovery of about a month at the impact site, relative to reference sites. In the south, MPB biomass did not change significantly as a result of dredging in the Longmore (2006b) study, but concurrent measurements by Edmunds et al. (2005) indicated an area-wide decline in biomass, unrelated to dredging. Though there was some evidence of increased DIN fluxes in the south during dredging, the reduction in denitrification

SEES Head Technical Report: Nutrient Cycling 70

efficiency was not statistically significant. This may again be due to a lower than desirable degree of replication, but any changes in denitrification as a result of changes in MPB productivity appear to be subtle, rather than dramatic. In the north, an almost complete loss of photosynthetic production was accompanied by a decline in denitrification efficiency of only 12%. Whether this would be true after a longer period of dredging impact is still unknown. However, indirect support for the view that MPB plays only a small role, if any, in denitrification is given by the following. Edmunds et al . (2005) documented seasonal declines in MPB biomass in high-biomass patches in both north and south from 50- 60 mg m -2 in summer, to less than 20 mg m -2 in winter. Though we do not have seasonal benthic flux measurements from the same sites, if this is a general pattern, and MPB played a significant role in denitrification, we could expect denitrification efficiency to be higher in summer than in winter. What evidence we have indicates the opposite: where we have summer and winter measurements (from 1994- 95), efficiency was higher in winter than summer. Despite uncertainty about the extent of impact on denitrification during trial dredging, it appears that the microbial system on which denitrification depends had recovered fully in both the north and south within a month of the end of dredging.

7.04 SEES sediment nutrient bioavailablity experiment

Though results of this work (Longmore 2006c) have been outlined in the appropriate sections above, details are brought together here to allow comparisons between areas. Thirty-eight sediment cores 90 mm in diameter, and to the full depth of expected dredging, were collected from the Yarra River/Williamstown Channel (11), Port Melbourne Channel (9), South Channel (10), South-east DMG (3) and the extension south of the Port Melbourne DMG (5). Cores were cut into 100 mm sections, and up to five sub-samples collected from each core to ensure that the major sediment types in each core were analysed chemically. One hundred and fifty-five sub-samples were elutriated with seawater (4:1 mixture by volume, for 4 hours), and the elutriate analysed for ammonium, oxidised N, dissolved organic N (DON), phosphate and silicate. These represent the nutrients immediately available for plant growth. The same sediment sub-samples were analysed for total N, total P and biogenic Si concentration, to allow an estimate of the absolute amount of nutrient likely to be disturbed during dredging. On the basis of the total N, P and Si analyses, 20% of the sub-samples were incubated in aerated seawater for 10 days, to measure the inorganic (bioavailable) nutrients released to solution over time. Longmore (2006d) estimated that all sediment particles > 4 µm in diameter should settle through a 10 m water column within eight days. Since less than 5% of the sediment in the north, and 1% in the south, is finer than 4 µm, 10 days was chosen as a reasonably conservative estimate of time in which sediment may remain suspended. Ammonium concentration in the elutriates varied by region and depth (Fig 41). The highest concentrations (almost 2,500 µM) were found at three depths between 1-4 m in one core (VB139) opposite the Williamstown dockyard. All other ammonium concentrations were less than 1,000 µM. Ammonium concentrations were low and varied little with depth in the South Channel and Port Melbourne Channel, and (apart from the high measurements from site VB139) varied little with depth in the Yarra River and Williamstown Channel. In contrast, ammonium concentrations increased markedly with depth for those samples from the extension south of the Port Melbourne DMG, from near zero at the surface, to ~ 1,000 µM at 5 m. Oxidised N (nitrate plus nitrite) concentrations were less than 0.2 µM for all but three samples, and were similar to the ambient level in the water column. No variation with depth was observed. DON concentration was similar to the ammonium concentration in the surface metre in all areas but much less (5-30%) at greater depths. According to these estimates, ammonium is the only nutrient of concern. Even if it were all bioavailable, DON is a much smaller load than ammonium. The major contribution to the ammonium load comes from the deep sediments in the Williamstown Channel, but the confidence interval is wide, because of a few very high concentrations at depth skewing the distribution of measurements. Sediment total N concentrations varied between 0-3,000 µg N g -1. Concentrations varied over a relatively narrow range in the PoM DMG extension (500-1,500 µg N g -1), with a slight decline with depth. Concentrations of up to 1,300 µg N g -1 were found in near-surface sediments in the Port Melbourne

SEES Head Technical report: Nutrient Cycling 71

Channel, but varied over a wide range at greater depths. Total N concentrations varied over the range 0- 1,000 µg N g -1 in the South Channel, and may have declined slightly with depth. There was no strong pattern with depth in the Yarra/Williamstown Channel cores, but the three highest values (to 2,200 µg N g-1) occurred in the surface 30 cm.

Ammonium ( µµµM) 0 500 1000 1500 2000 2500 3000 0 1 2 3 South Ch 4 Depth (m) Depth 5 6

Ammonium ( µµµM) 0 500 1000 1500 2000 2500 3000 0 1 2 3 PM Channel 4 Depth (m) Depth 5 6

Ammonium ( µµµM) 0 500 1000 1500 2000 2500 3000 0 1 2 3 Yarra/Wtown 4 Depth (m) Depth 5 6

Figure 41 Ammonium concentrations in elutriates from each area (Longmore 2006c).

SEES Head Technical Report: Nutrient Cycling 72

Yarra River Williamstown Channel

35 300

30 250

25 VB106 0.1-0.2m 200 sediment) sediment ) sediment VB139 0.1-0.2m -3 20 VB113 0.1-0.2m -3 VB139 0.4-0.5m VB113 0.5-0.6m 150 VB139 2.4-2.5m 15 VB124 1.5-1.6m VB139 3.6-3.7m VB125 0.1-0.2m 100 10

50 Ammonium (g N m N (g Ammonium 5 m N (g Ammonium

0 0 0 2 4 6 8 10 12 0 2 4 6 8 10 12 Days elapsed Days elapsed

Port Melbourne Channel Blanks & Borrow ground

45 8

40 7 35 6 30 VB145 0.1-0.2m

sediment) sediment) 5 Blank 1 -3 VB146 0.1-0.2m -3 25 Blank 2 VB146 0.4-0.5m 4 20 BAS25 0.4-0.5m VB147 0.1-0.2m 3 BAS25 1.4-1.5m 15 VB152 0.2-0.3m 2 10

Ammonium (g N m N (g Ammonium 5 m N (g Ammonium 1

0 0 0 2 4 6 8 10 12 0 2 4 6 8 10 12 Days elapsed Days elapsed

South Channel SE DMG

1.2 7

1.0 6

5 VB27 0.4-0.5m 0.8 VB14 0.7-0.8m sediment)

sediment) VB27 4.4-4.5m -3

VB17 0.4-0.5m -3 4 VB28 0.5-0.6m 0.6 VB17 1.4-1.5m VB28 2.4-2.5m VB23 2.4-2.5m 3 VB29 1.2-1.3m 0.4 VB26 2.8-2.9m 2 VB29 2.4-2.5m 0.2 Ammonium (g N m N (g Ammonium

Ammonium (g N m N (g Ammonium 1

0.0 0 0 2 4 6 8 10 12 0 2 4 6 8 10 12 Days elapsed Days elapsed

Figure 42 Release of ammonium from sediment particles over 10 days, after suspension in seawater (Longmore 2006c). In general, ammonium, oxidised N and silicate were released from the sediment particles over 10 days, while DON and phosphate changed little (Fig 42). However, sediments from the South Channel took up ammonium, rather than released it. The source of the nutrients that became bioavailable over 10 days was reasonably evenly divided between the Williamstown Channel, Yarra River and Port Melbourne Channel for ammonium and silicate. Surficial sediments were the main ammonium and silicate sources in the Port Melbourne Channel and Yarra River, while deeper sediments dominated for the Williamstown Channel. Surficial sediments in the Port Melbourne Channel and the Yarra River contributed most to the phosphate release. The fact that ammonium and/or oxidised N increases did not relate to decreases in DON suggests that the source of the bioavailable N is release from particles, rather than breakdown of DON. 7.05 Estimates of nutrient loads from dredging

When the estimated loads to the Bay from each of these nutrient forms is calculated from predicted dredge volumes (Longmore 2006c), the vast majority (> 98%) of the N, P and Si load in the sediments to be dredged can be considered inert. However, small quantities of ammonium (~ 50 t) will be released immediately at the dredge site, and small quantities of ammonium (~ 75 t) will also be released from the dredge plume and areas in which it settles over 10 days. The DON load is about one-third of the dissolved inorganic (almost entirely ammonium) load. The confidence intervals around each mean are

SEES Head Technical report: Nutrient Cycling 73

large, but the combined bioavailable N load (mean ~ 125 t, or 332 t at the 95 th percentile) is lower than that estimated for the EES. The study examined the total nutrient content (N, P and biogenic Si) in sediments likely to be mobilised by dredging (Table 16). This indicated that dredging in the South Channel, followed by the Williamstown Channel, will contribute most to the total N and P loads mobilised by dredging. The trailing suction hopper dredge works by sucking up sediment in a stream of water, during which process the nutrients held in sediment pore water, or very weakly bound to particle surfaces, may be released into the water phase. The study also examined bioavailable nutrient release into the dredge water, which would either be retained in the hopper (in non-overflow mode) or pumped back into the water column at the dredge site (in overflow mode). This impact is estimated in Table 17. Dredging in the Williamstown Channel is likely to lead to the largest input of labile N from the dredge water, with the total input from all dredging areas substantially lower than estimated for the EES. Lastly, the study examined the longer-term release of nutrients from suspended particles, both while they are still in the water column, and after they have settled to the seafloor (Table 18). These three aspects of nutrient release from the sediment set an upper level to the likely impact of one aspect of dredging on nutrient cycling.

Table 16 Estimated total nutrient load (mean ±±± 95% CI) mobilised by dredging in each area (Longmore 2006c)

Area Total N Total P Biogenic Si (tonnes N) (tonnes P) (tonnes Si) Yarra River 2,200 ± 6,250 975 ± 600 16,900 ± 2,900 Williamstown Channel 1,050 ± 1,450 460 ± 380 7,200 ± 5,600 Port Melbourne Channel 740 ± 300 450 ± 240 8,300 ± 2,200 South Channel 8,000 ± 5,100 4,350 ± 1,800 17,200 ± 9,100

Total 12,000 ± 13,100 6,250 ± 3,000 49,600 ± 19,800

Table 17 Estimates of dissolved nutrient load ( mean ± 95% CI) released to dredge water (Longmore 2006c).

Area Ammonium Oxidised N DON Phosphate Silicate (tonnes N) (tonnes N) (tonnes N) (tonnes P) (tonnes Si)

Yarra River 16.4 ± 23.9 0.03 ± 0.02 10.0 ± 11.1 1.0 ± 0.7 19.8 ± 5.5 Williamstown 17.7 ± 71.0 0.01 ± 0.01 5.8 ± 9.0 0.8 ± 1.5 5.6 ± 3.7 Channel Port 3.1 ± 2.8 0.02 ± 0.01 4.3 ± 3.4 0.9 ± 0.6 7.0 ± 2.5 Melbourne Channel South Channel 9.6 ± 6.2 0.08 ± 0.00 11.3 ± 2.7 1.1 ± 0.7 29.8 ± 6.2 Total 46.8 ± 105.5 0.15 ± 0.04 31.3 ± 26.2 3.8 ± 3.8 62.0 ± 17.9

SEES Head Technical Report: Nutrient Cycling 74

Table 18 Estimates of bioavailable nutrient load ( mean ± 95% CI) released from particles after settling (Longmore 2006c).

Area Ammonium Oxidised N DON Phosphate Silicate (tonnes N) (tonnes N) (tonnes N) (tonnes P) (tonnes Si)

Yarra River 20.8 ± 14.4 0.7 ± 0.7 -1.8 ± 3.0 1.3 ± 1.8 14.8 ± 5.5 Williamstown 23.8 ± 20.0 0.2 ± 0.2 1.0 ± 2.3 0.6 ± 0.9 7.7 ± 3.6 Channel Port 31.5 ± 40.8 1.8 ± 2.4 1.2 ± 2.3 2.0 ± 2.0 20.1 ± 19.1 Melbourne Channel South Channel -2.6 ± 24.9 4.6 ± 15.5 8.5 ± 10.4 0.2 ± 2.2 22.6 ± 13.8

Total 73.5 ± 101 7.3 ± 18.8 8.8 ± 18.1 4.1 ± 6.9 65.2 ± 42.0

The bioavailable (ammonium plus oxidised N) N load from both the pore water (released in the dredge hopper) and from particles (released up to 10 days after suspension) is therefore 130 ± 226 tonnes. This is compared to some of the known or estimated DIN inputs or pools in the Bay (Table 19).

Table 19 Estimates of bioavailable N load from dredging compared to known existing inputs, fluxes and pools (Harris et al . 1996; *Murray and Parslow 1999; Longmore 2006c). Source DIN Inputs: Immediately released (Table 17), tonnes 47 ± 106 Released over 10 days (Table 18), tonnes 83 ± 120 WTP DIN input (Fig 22), t y -1 2,500 - 3,500 River DIN, t y -1 740 – 2,500 Fluxes: Filter-feeder excretion, t y -1* 4,300 Water column recycling, t y -1* 19,200 Benthic flux, t y -1* 6,200 Denitrification, t y -1* 6,900

(as N 2 rather than DIN) Pools: Water column DIN, t 230 - 400 Phytoplankton, t* 210

SEES Head Technical report: Nutrient Cycling 75

The total N annual load to the Bay is thought to vary by about 3,000 t, and the assumption in the PPBES model is that almost all of it eventually becomes bioavailable (Parslow et al. 1999). The additional bioavailable N load likely to be generated by dredging is well within interannual variation in the “natural” load. At the 95 th percentile of the mean, it is of the same order as the N load delivered during the 1993 flood in the Yarra (~ 500t; Harris et al . 1996, Fig 3.3), but will be delivered over a much longer period. Another way of estimating the impact of release of nutrients during dredging is to calculate the impact per unit area. Dredging of contaminated material in the Yarra River is planned in non-overflow mode, while all other dredging is in overflow mode. In principle, most of the labile nutrients from the sediments dredged in non-overflow mode will be transported to the PoM DMG, while those dredged in overflow mode will disperse from the dredging area, over at least the area covered by the plume. Modelling of the dredge plume is used to indicate the area of impact (Figs 43-45). The 2 mg L -1 plume contour is chosen here because it indicates the sphere of influence of the finest particles, which are likely to be highest in nutrient content and to travel the greatest distance before settling (Longmore 2006a). The outer contour, which applies for the least proportion of time (<1%), is also chosen as a very conservative indication of the area influenced by the plume. The model indicates that the impact of the plume from dredging contaminated sediments in the Yarra will be confined to the river and most of Hobsons Bay. This is because dredging of this material will be carried out in non-overflow. In contrast, uncontaminated material in the river and Williamstown Channel, which will be dredged in overflow mode, both create an impact well outside Hobsons Bay (Fig 43). The impact from dredging in the Port Melbourne Channel (Fig 44) occupies most of the area east and north of the channel to the shore. Dredging in the South Channel (Fig 45) affects all of the area on and south of the Sands, and some of the deeper water north of the Sands.

SEES Head Technical Report: Nutrient Cycling 76

Figure 43 Contours for the fraction of time suspended solids exceed 2 mg L -1 in the dredge plume during dredging of contaminated (top LHS) and uncontaminated (top RHS) sediment in the Yarra River and Williamstown Channel (bottom RHS); CLT 2006.

SEES Head Technical report: Nutrient Cycling 77

Figure 44 Contours for the fraction of time suspended solids exceed 2 mg L -1 in the dredge plume during dredging in the Port Melbourne Channel (CLT 2006).

SEES Head Technical Report: Nutrient Cycling 78

Figure 45 Contours for the fraction of time suspended solids exceed 2 mg L -1 in the dredge plume during dredging in the South Channel east of Hovell Pile by the Cornelis Zanen (top LHS) and the Queen (top RHS), west of Hovell Pile (bottom LHS) and in South Channel West (bottom RHS); CLT 2006.

SEES Head Technical report: Nutrient Cycling 79

From Fig 43, the area of the plume from dredging of contaminated material in the Yarra is about 12 km 2, while that for dredging uncontaminated sediment in the Yarra is about 42 km 2, and that for sediment from the Williamstown Channel is 50 km 2. The area affected at the PoM DMG is about 20 km 2. The area of the plume for dredging in the Port Melbourne Channel is about 150 km 2 (Fig 44), while the area under the plume for the three dredging areas in the South Channel (Fig 45) is about 600 km 2 (inside the Bay, including the SE DMG). Using the periods of dredging shown in the figures, and applying the immediately released bioavailable N estimates for each dredging area, it is possible to calculate the N loading per square metre per day attributed to each activity (Table 20). The assumptions involved include: • in overflow dredging, at a plume discharge rate of 200 kg s -1 for a jumbo TSHD, and 90 kg s -1 for a mid-sized TSHD (CLT 2006), about 2 % of the dredged sediment is lost in the plume, and the rest is placed at the PoM DMG, while 100% of the dissolved nutrient is lost in the plume; • in non-overflow dredging, 1% of the dredged sediment enters the plume, 99% remains in the hopper and 1% of the dissolved nutrient enters the plume; • 50% of the Yarra sediments are dredged in non-overflow, and 50% in overflow; • slowly released bioavailable N is divided in the same proportions as the dredged sediment between plume and PoM DMG; • even if buried by subsequent loads, bioavailable N will be released from all placed sediment. These estimates will be conservative, because the last dot point is clearly not correct, and also because the slowly-released bioavailable N estimates were from those sediments with highest nutrient content (and so probably over-estimate bulk sediment nutrient concentrations). As a yardstick for these estimated loading rates, the CSIRO model (Murray and Parslow 1999) predicted annual mean sediment DIN fluxes of > 20 mg N m -2 d -1 in Hobsons Bay, 10-20 mg N m -2 d -1 near the PoM DMG, 5-15 mg N m -2 d -1 in the north, and < 10 mg N m -2 d -1 just north of, on and south of the Sands. Recent benthic flux monitoring (for the SEES) indicated DIN fluxes from the sediment of 30-80 mg N m -2 d-1 in Hobsons Bay, 2-7 mg N m -2 d -1 near the Port Melbourne Channel, 1-6 mg N m -2 d -1 on the PoMDMG, 3-10 mg N m -2 d -1 on the SEDMG, and 0-6 mg N m -2 d -1 on the Sands. Except on the DMGs, the additional dredging load from immediately-released and slowly-released bioavailable N is much less than the existing sediment DIN release. Dredging will lead to an order of magnitude increase in DIN loading at the PoM DMG over current levels, though much of this increase is from release assigned to particles that may actually be buried rather than exposed on the surface. This information is summarised in Table 21.

SEES Head Technical Report: Nutrient Cycling 80

Table 20 Estimates of nutrient loading from each dredging activity (Longmore 2006c).

Dredging activity Area affected Area of plume Dredging Loading rate (km 2) duration (weeks) (mg N m -2 d -1), mean and 95% CI Immediately bioavailable N Contaminated Yarra/Hobsons 12 4 0.1-0.5 Yarra sediments Bay Contaminated PoMDMG 20 4 8-30 Yarra sediments Uncontaminated Yarra/Hobsons 42 8 2-7 Yarra sediments Bay Williamstown Yarra/Hobsons 50 18 1-12 Channel sediments Bay Port Melbourne North 150 16 0.1-0.3 Channel sediments South Channel South 600 40 0.05-0.08 sediments

Slowly-released bioavailable N Contaminated Yarra/Hobsons 12 4 0.3-0.5 Yarra sediments Bay Contaminated PoMDMG 20 4 19-32 Yarra sediments Uncontaminated Yarra/Hobsons 42 8 0.1-0.2 Yarra sediments Bay Uncontaminated PoMDMG 20 8 9-16 Yarra sediments Williamstown Yarra/Hobsons 50 18 0.1-0.2 Channel sediments Bay Williamstown POMDMG 20 18 9-17 Channel sediments Port Melbourne North 150 16 0-0.1 Channel sediments Port Melbourne POMDMG 20 16 14-33 Channel sediments South Channel South 600 40 0-0.01 sediments South Channel SEDMG 20 40 0.4-7.4 sediments

SEES Head Technical report: Nutrient Cycling 81

Table 21 Existing benthic DIN fluxes compared to estimated DIN release from porewater and particles. Dredging loadings are means and 95 th percentile (Longmore 2006c).

Area Existing benthic DIN Estimated immediate Estimated slow DIN flux range DIN release release (mg N m -2 d -1) (mg N m -2 d -1) (mg N m -2 d -1) Hobsons Bay 30-80 6-39 <1 Port Melbourne 2-7 0.1-0.3 <0.1 Channel Port Melbourne DMG 1-6 14-39 39-112 South east DMG 3-10 - 0.4-7 The Sands 0-6 <0.1 <0.1

Further modelling to estimate flood loads to the Bay (Parslow et al. 2000) is also illuminating. The PPBES model was run to assess the risk to the Bay of a rare event, a 1 in 25 year flood in the Yarra. In fact, three hypothetical floods were modelled. The hypothetical floods discharged 190, 219 and 326 t DIN to Hobsons Bay over about 30 days. All of the impacts were restricted to Hobsons Bay and the east coast; the floods had little impact on the central zone of the Bay. Under the lowest flood input, mean chlorophyll in Hobsons Bay declined, because of increased light attenuation. Under the highest input, mean chlorophyll increased by about 60% along the east coast (and less in Hobsons Bay). Under the two highest loads, the model predicted median denitrification efficiency and microphytobenthos productivity in Hobsons Bay of zero. The model showed persistent salinity stratification during the flood event, with draw-down of bottom-water oxygen concentrations expected to be even larger than the model prediction of anoxia in bottom waters over 10-20 days. The smallest modelled flood load is 50% larger than the mean load (immediately- and slowly-released DIN) expected for the whole dredging program (Table 18). The largest modelled flood load is similar to the 95%ile estimate of the dredge load, but delivered over a much shorter period.

SEES Head Technical Report: Nutrient Cycling 82

7.1 Risk Assessment Method Introduction A detailed Project risk assessment has been conducted as part of the SEES process to evaluate the potential impacts that the Project will or could have on a wide range of assets, values and beneficial uses. This study has contributed to this risk assessment and the results have been used to help form the conclusions of this study. The risk assessment methodology and implementation has been guided by URS as Risk Advisor to PoMC. A detailed description of the process is provided in the URS Risk Technical Report. This section of the report provides a brief summary of the risk assessment process and how technical specialists have been involved in the process. Risk Assessment Process The key aspects of the process are summarised as follows. For a detailed description of the risk assessment process, please refer to the URS Risk Technical Report. • The risk assessment was conducted for the purposes of the SEES and considers potential environmental, social and economic impacts of the Project on the wider environment and community. • The risk assessment is based on the Project Description and its outputs represent the potential impacts of implementing the Project as described in the Project Description. • The risk assessment process was based on an event tree approach. The event trees, which describe clear linkages between cause and effects, formed the basis of the risk assessment and the framework for identification of likelihoods and consequences in relation to the assets discussed in this report. The pathways described in the event trees were constructed based on substantial consultation, advice and review by relevant technical specialists. • The risk assessment process used a Likelihood Guide and Consequences Table to ensure a consistent assessment of likelihood and consequence across all specialist areas within the Project. The Likelihood Guide used, shown in Table 22, is from a published source (URS 2006). The Consequences Table used for estimating diverse consequence types on an even basis was developed specifically for the Project based on substantial consultation and advice from the technical specialists. The categories of the Consequences Table relevant to this study are shown in Table 23. Algal blooms, nutrient cycling and denitrification are all classed as ecosystem functions. The Risk Consultants provided a scale of consequences for ecosystem-level impacts, which varied by an order of magnitude for each level. Consequence Levels vary in relation to natural variability, change in ecosystem function, and period of recovery.

Table 22 Likelihood Guide

Qualitative Description Order of Magnitude Basis Annual Probability A. Certain 1 (or 0.999, 99.9%) Certain, or as near to as makes no difference B. Almost certain 0.2-0.9 One or more incidents of a similar nature has occurred here C. Highly probable 0.1 A previous incident of a similar nature has occurred here D. Possible 0.01 Could have occurred already without intervention E. Unlikely 0.001 Recorded recently elsewhere F. Very unlikely 0.0001 It has happened elsewhere G. Highly improbable 0.00001 Published information exists, but in a slightly different context H. Almost impossible 0.000001 No published information on a similar case

SEES Head Technical report: Nutrient Cycling 83

Table 23 Environmental Consequences Table

Negligible Minor Moderate Major Extreme

Minimal, if any Low level impact High level of High level of High level of impact for some for some impact for some impact for impact State- communities. communities, or communities, or communities bay- wide. CONSEQUENCE Potentially some high impact for a moderate impact wide. LEVEL impact for a small small number for communities number (<10) of (<10) of bay-wide. individuals. individuals.

0.1 1 10 100 1000

Ecosystem Alteration or Measurable Measurable Measurable Long term and Function (including disturbance to changes to the changes to the changes to the possibly resilience and ecosystem within ecosystem ecosystem ecosystem irreversible resistance) natural variability. components components components with damage to one or Ecosystem without a major without a major a major change more ecosystem interactions may change in change in in function. function. have changed function (no loss function (no loss Recovery (ie Recovery, if at but it is unlikely of components or of components or within historic all, greater than that there would introduction of introduction of natural variability) 10 years be any new species that new species that in 3 to 10 years following detectable affects affects following completion of change outside ecosystem ecosystem completion of Project natural variation / function). function). Project construction. occurrence. Recovery in less Recovery in 1 to construction. than 1 year. 2 years following completion of Project construction.

Habitat, Alteration or 1 to 5% of the 5 to 30% of the 30 to 90% of the Greater than communities and / disturbance to area of habitat area of habitat area of habitat 90% of the area or assemblages habitat within affected in a affected in a affected in a of habitat natural variability. major way or major way or major way or affected in a Less than 1% of removed. Re- removed. Re- removed. Re- major way or the area of establishment in establishment in establishment in removed. Re- habitat affected less than 1 year 1 to 2 years 3 to 10 years establishment, if or removed. (relative to following following at all, greater component completion of completion of than 10 years seasonality) Project Project following following construction. construction. completion of completion of Project Project construction. construction.

• Technical specialists from all of the specialist areas relevant to the Project were involved in all aspects of the risk assessment process. All of the risk assessment inputs including likelihood and consequence ratings were provided by technical specialists, and assessments were discussed in a number of workshops. The assessment of risks to nutrient cycling drew most heavily on advice from the plume modelling, marine ecology, sediment chemistry and water quality studies.

SEES Head Technical Report: Nutrient Cycling 84

• The input data for the risk assessment was obtained through a process that involved substantial consultation and interaction within and between different technical specialist disciplines. This consisted of numerous workshops, meetings, discussions and data reviews over a period of several months. • The risk assessment outputs have been derived by applying a systematic approach to develop estimates of risk and to develop profiles for each asset showing predicted effects (those that are certain or almost certain to occur) and risk events (those events and effects that could, but are not expected to occur) relating to the asset. (Refer to the URS Risk Technical Report for more detail). Predicted effects are discussed in terms of the Consequence Level assigned.

• The risk (expressed as the “risk quotient”) was calculated as the likelihood of occurrence of an event over the Project, multiplied by the magnitude of its consequences. (Refer to the URS Risk Technical Report for more detail). Risk outputs of relevance to this report are summarised and discussed in the following sections. The risk quotient was assessed as “major”, “moderate”, “minor”, or “very low”, which vary from one to the next by a factor of 10. • Note: The risk assessment process has been designed to assist in understanding the predicted effects and risks of the Project. It is based on professional judgments and deals with complex concepts of a dissimilar nature in an environment of uncertainty. For further description of the treatment of uncertainty in the risk assessment process, please refer to the URS Risk Technical Report. Uncertainty has been incorporated into the assessment in the sense that likelihood levels and consequence levels differ by an order of magnitude. Unless the likelihood or consequence of an impact is uncertain to more than a factor of 10, the descriptive risk quotient will not be changed. Some of the risks to nutrient cycling have a seasonal sensitivity, and they are best assessed against the indicative dredging schedule (Table 24), outlined in the CDP Dredging Strategy (PoMC Sep 2006). This strategy assumes a January 2008 start. Dredging is planned in full overflow mode, except for contaminated sediments in the Yarra and Williamstown Channel, which will be dredged in non-overflow mode in May-June 2008.

Table 24 Indicative dredging schedule for each project area.

Area Dredging period(s) Sediment type Total dredging time (weeks) North (Port Melbourne Jan-May 2008; Dec Stiff clays 16 Channel) 08-Jan 09. South (South Channel) Mar-May 08; Feb-Sep Silty sands 48 09. Yarra/ Hobsons Bay May 08; June-Aug 08; Soft silts, harder clays 26 (Williamstown Channel) beneath Feb-May 09. Yarra/ Hobsons Bay Jun 08; Dec 08-Jan 09. Soft silts, harder clays 13 (Yarra River) beneath South (The Entrance) Jun 08-Jan 09 Weak rock 27

The activities are further divided into Predicted events-those for which the likelihood of occurrence is 0.5 or higher (ie. the event is certain or almost certain to occur), and Risk events, which have a likelihood of less than 0.5.

SEES Head Technical report: Nutrient Cycling 85

7.2 Bay-wide

No event, either predicted or risk, is expected to lead to a risk quotient of medium or higher for nutrient cycling Bay-wide. Predicted events

The issue of a change in tidal flushing, as a result of channel deepening, leading to a change in nutrient flushing, has been classed as a predicted event, because both are almost certain to occur. However, given the relatively small changes anticipated (Fig 6), and the long residence time relative to phytoplankton generation times, even in the south, this is thought to have an insignificant consequence to nutrient cycling on both regional and bay-wide scales, and the risk quotient is very low.

Risk events

Only one issue has been defined which may have a Baywide impact on nutrient cycling. This is the introduction or redistribution of a marine pest to the Bay, by the arrival of dredging and/or monitoring vessels, or their subsequent use in the Bay (Table 25). Given the efforts that will be made to clean and inspect the hulls of all vessels prior to arrival in the Bay for dredging, the likelihood of a new pest being introduced is almost impossible. The likelihood of the pest, once introduced, spreading to most of the Bay, and affecting denitrification, is very unlikely (for each of the four project areas). As noted above, studies (Ross et al . unpublished) have indicated that most marine pests have little long-term impact on the ecosystem; after an initial “boom”, the population settles to a lower level. Those introduced pests in the Bay found to affect nutrient cycling, and denitrification in particular, only do so at densities much higher than found in the Bay. The total likelihood is therefore still almost impossible. However, if a pest were to spread Baywide, and affect denitrification, the consequence would be extreme. Despite this, the risk quotient is very low.

All of the other potential impacts described in Section 6 may have localised impacts (see below), but none are expected to lead to Bay-wide impacts on nutrient cycling.

SEES Head Technical Report: Nutrient Cycling 86

Table 25 Summary of bay-wide risks to nutrient cycling. Predicted effect Aspect Activity Initiating Impact A Impact B Consequen event ce level Dredging Deepening of Change in Change in nutrient Change in Very low TSHD channel hydrodynamic flushing nutrient processes cycling Risk events Aspect Activity Initiating Impact A Impact B Consequen Risk event ce level Quotient Vessel Vessel use Transfer Translocation of Denitrification Major Very low Management and between marine pests change due to management project areas altered fauna Vessel Survey and Transit of Translocation of Denitrification Major Very low Management monitoring survey and marine pests change due to monitoring altered fauna vessel (changes to fauna) Vessel Survey and Transit of Translocation of Denitrification Major Very low Management monitoring survey and marine pests change due to monitoring altered fauna vessel (MPB) Vessel Vessel use Vessel Vessel / Introduction of Extreme Very low Management and mobilisation equipment has marine pests management marine pests new to the Bay

SEES Head Technical report: Nutrient Cycling 87

7.2 Yarra River and Hobsons Bay The highest-rated risk to nutrient cycling in this area, release of labile nutrients from the dredge, is low. Dredging activities in this area include TSHD dredging of the Yarra River and Williamstown Channel, and backhoe dredging of berths and channel batters. Dredging of contaminated sediment will be carried out in non-overflow mode by a jumbo-sized TSHD in autumn; all other material will be dredged in overflow mode by a mid-sized TSHD in summer, autumn and winter. This area of the Bay is already subject to high nutrient loads, algal blooms, both toxic and non-toxic, and reduced denitrification efficiency. The PPBES model predicts low denitrification efficiency in the Yarra mouth (mean < 50%, and zero 6% of the time), and measured efficiency in Hobsons Bay is around 50%. Activities assessed are listed in Table 26. Predicted events Physical removal of the seabed is certain; subsequent loss of an effective denitrification layer is less certain but possible, since the PPBES model assigned low efficiency to the Yarra. The consequence is considered minor, because the channel bed comprises only a small proportion of the seabed, and may be less effective at denitrifying than undisturbed sediments in Hobsons Bay. Even if there is a change, it will not be major, and should recover in less than a year. In this area, dredging has been a frequent event, and the infauna is most likely dominated by rapidly-colonising species. The risk quotient is very low. Turbidity during dredging is likely to be so high in the Yarra River (Fig 43) that reductions in phytoplankton production in the river are certain, despite the expected nutrient enrichment (see Appendix 1). Given the small area of the Yarra compared to Hobsons Bay, and the series of relatively short impacts (totalling 13 weeks in the Yarra, 26 weeks in the Williamstown Channel), and the likely rapid increase in light once dredging ceases, the loss in production is thought to be minor. This leads to a medium risk quotient. Hobsons Bay is subject to occasional high turbidity from natural causes, over a background of about 5-10 mg L -1 (Longmore et al . 1996; You et al. 1996). Phytoplankton production is high in Hobsons Bay, particularly in summer. The plume modelling (not shown) indicates that a TSS increase of < 5 mg L -1 is expected over most of Hobsons Bay for dredging in the Williamstown Channel (26 weeks), with lower increases expected for dredging in the Yarra. The expected increase is therefore within natural variation for Hobsons Bay, but for a longer period than usual. The impact on phytoplankton production is likely to be higher during calm periods than during windy periods, when turbulent mixing should expose cells to light for at least part of the day. As Hale (2006b) points out (Appendix 1), possible outcomes in Hobsons Bay range from turbidity suppression of production to nutrient enhancement, with a range of responses between the two likely. The net impact on denitrification is uncertain, partly because high oxygen demand at the sediment surface is one reason for the low denitrification efficiency observed in Hobsons Bay. The impact could, in theory, be beneficial, if a reduction in phytoplankton production led to an increase in oxygen concentration at the sediment surface. Even assuming a negative impact on denitrification is possible, given that recovery is expected within weeks to months of the end of dredging, the consequence is minor. The risk quotient is very low. Turbidity may also impact on MPB production. Since MPB require light to photosynthesise, production should decline under the dredge plume. MPB productivity is about three times higher in summer than winter (Beardall and Light 1996), so the impact may be more severe in summer. MPB biomass is not particularly high in Hobsons Bay (Edmunds et al . 2006). The experimental evidence we have from the Trial Dredge program (from the Port Melbourne Channel, rather than Hobsons Bay) does not support the likelihood of a decline in MPB production leading to a major reduction in denitrification. However, dredging was brief during the TDP, compared to that planned for the CDP. For the sake of argument, let us assume that MPB contribute 50% toward denitrification of 3.1 mg N m -2 d -1 (Table 6), which is lost completely from the whole of Hobsons Bay for 30 weeks (26 weeks of dredging plus one month recovery). The impact of denitrification loss due to shading of MPB is the equivalent of an additional N load to Hobsons Bay of about 20 mg N m -2 d -1. This is similar to the current benthic DIN flux (Table 21), and within the range of natural variation. On this basis, the likelihood of a loss in MPB productivity is possible, and the consequence to nutrient cycling is minor. The risk quotient is very low.

SEES Head Technical Report: Nutrient Cycling 88

Risk events The dredged material contains labile nutrients, which will be released in overflow, or at the PoM DMG (Table 20). However, given the size of the loading, the period of dredging in this area (summer through to winter), and the annual load already carried by the Yarra (~ 700-2,200 t N), the dredging load does not amount to a dramatic increase over background in this area (Table 21). At very worst, the DIN loading per unit area in Hobsons Bay will double, but most likely it will increase by 10%. The production of labile nutrients is certain, but the likelihood that this leads to an increase in plankton production and deposition, and then in a decline in denitrification is unlikely to possible. If it did, the consequence is minor, because the change may well be within natural variation, with recovery within a year. Note that the estimated load is substantially lower than that previously estimated for the EES. The risk quotient is low. Similarly, the dredged material contains nutrients bound to particles, some of which may be released over time as particles settle. The probable area of impact is indicated by model output of the dredge plume (Fig 43); the issue to be resolved is how significant the load is. Assuming the plume affects the whole of the project area (from the plume modelling), the particle-derived bioavailable N load is less than 3% of the current benthic DIN flux in that area. The likelihood that labile nutrients will be released is almost certain, but the probability that this will lead to a decline in denitrification is unlikely, simply from the small increase over background. The consequence is minor, again because the change may be within natural variation. The risk quotient is very low. The TSHD dredge plume will mobilise algal cysts (this has already been observed in maintenance dredging, though the most recent study failed to find toxic algal cysts in the sediments of the Yarra and Hobsons Bay- SKM 2004). There is no certainty that this will lead to an algal bloom, but such an outcome is possible, and more likely in summer-autumn than other seasons. The main impacts are: (i) economic (transfer of toxic bloom to aquaculture site) but the only site adjacent is not in commercial production, and (ii) ecological, on benthic organisms and nutrient cycling, if the bloom is large enough to consume bottom DO on settling, or to add substantially to the carbon load to the sediment. We estimate a 30% chance of that happening, so that the net likelihood is between unlikely and possible. To be conservative, the higher likelihood is taken. The consequence to nutrient cycling is minor, because of the restricted period of impact. For example, denitrification efficiency may be reduced in the Yarra River and/or Hobsons Bay, but is expected to recover within a year. The risk quotient is very low. Economic and human health consequences are dealt with by others. Sedimentation from the dredge plume may also impact on MPB. There is evidence that some types of cells in the MPB (but not all) can migrate in the sediment to compensate for burial (Edmunds et al . 2004). At the worst, the impact of burial would be the same as that estimated for the impact of turbidity (complete loss of productivity over the whole of Hobsons Bay by one or other route); the only difference may be recovery time. Even if recovery took a year, for the same reasons listed above, with a likelihood of possible, and a consequence to nutrient cycling of insignificant, the risk quotient is very low. Sedimentation from the dredge plume may lead to the burial of infauna, and therefore of whatever contribution they make to denitrification. Modelling (not shown here) indicates heavy sedimentation throughout the Yarra River, and either side of the Williamstown Channel. However, for most of Hobsons Bay, the sedimentation is expected to be light (< 1250 g m -2). Sediment traps in Hobsons Bay measured deposition rates of about 8 g m -2 h -1 (You et al . 1996), suggesting that natural sedimentation over a period similar to the dredging (~3,500 g m -2) exceeds that expected from dredging. The existing fauna may therefore be adapted to such deposition rates, or comprise rapidly colonising species that recover rapidly from such impacts, and the extra from dredging may make little difference to faunal activity. On this basis, though sedimentation is certain, the probability that it leads to a loss in infauna is rated as unlikely, and the consequence as minor, with recovery expected within a year. The risk quotient is therefore very low. Toxicants may affect grazers on phytoplankton. However, the contaminated sediments will be dredged in non-overflow mode, implying that there will be no remobilisation at the dredge area. If there was an impact, it is more likely to be confined to the Yarra River, rather than the whole of Hobsons Bay. The uncontaminated sediment, to be dredged in overflow mode, has not been found to be toxic (URS 2006). The likelihood of the chain of events is therefore unlikely; the consequence is minor, and the risk quotient is very low.

SEES Head Technical report: Nutrient Cycling 89

Toxicants may affect infauna, but not in this area, since any toxic sediments will be transported to the PoM DMG. For the same reasons as listed above, the likelihood of the chain of events is therefore unlikely; the consequence is minor, and the risk quotient is very low. The increased frequency of toxic algal blooms because of reduced Si inputs from catchments, caused by dams, has been observed (Rocha et al . 2002). The phytoplankton in Hobsons Bay are often dominated by non-siliceous species, and it could be argued that dredging is likely to raise Si levels, increasing the likelihood of diatom blooms, but reducing the risk of dinoflagellate and other potentially toxic blooms. Such a “positive” outcome has been ignored in these estimates.

Table 26 Summary of risks to nutrient cycling in the Yarra River and Hobsons Bay.

Predicted effects Aspect Activity Initiating Impact A Impact B Consequence event level Dredging Dredging - Increased Reduced light Impacts Minor TSHD creation of suspended phytoplankton plume sediments photosynthesis Dredging Dredging - Increased Reduced light Impacts MPB Minor TSHD creation of suspended photosynthesis plume sediments Risk events Aspect Activity Initiating Impact A Impact B Conse- Risk event quence Quotient level Dredging Dredging - Increased Algal blooms Minor Very Low TSHD creation of bioavailable plume nutrients Dredging Dredging - Increased Increased Algal blooms Minor Very Low TSHD creation of suspended bioavailable plume sediments nutrients Dredging Dredging - Increased Reduction in Minor Very Low TSHD creation of bioavailable denitrification plume nutrients Dredging Dredging - Mobilisation Algal blooms Transfer to Minor Very Low TSHD creation of of algal cysts sensitive area plume Dredging Dredging - Physical Removal of Reduction in Minor Very Low removal of removal of denitrification denitrification seabed seabed layer Dredging Dredging - Increased Algal blooms Minor Very Low Backhoe creation of bioavailable and Grab plume nutrients Dredging Dredging - Mobilisation Algal blooms Affects public Moderate Very Low TSHD plume of algal cysts health & safety

SEES Head Technical Report: Nutrient Cycling 90

7.3 North of the Bay

The highest risks to nutrient cycling in this area are all very low. The eastern shore of this area is subject to sporadic high nutrient loads from the Yarra, and the western shore is impacted by the WTP discharge, while the deep central waters are much poorer in nutrients and phytoplankton. Denitrification is highly efficient in the centre of the Bay, and there is some evidence that efficiency is lower at some sites (but not all) in shallow waters closer to shore. Because of the size of this project area (almost half of the Bay surface) and the overall high efficiency, this area makes a major contribution to the denitrification capacity of the Bay. The dredging activities in this area include 16 weeks of dredging of the Williamstown Channel by a medium-sized dredge, construction of the bund at the PoM DMG, and placement of material from the Yarra River, Williamstown Channel and Port Melbourne Channel. Activities evaluated are summarised in Table 27. Predicted events Physical removal of the seabed is certain; subsequent loss of an effective denitrification layer is less certain but probable, since the sediments in the channel appear to be organic-poor, and therefore presumed to play little role in denitrification. The consequence is considered negligible, because the channel bed comprises only a small proportion of the seabed, and may be less effective than undisturbed sediments. The risk quotient is very low. Risk events The dredged material contains labile nutrients, which will be released in overflow, or at the PoM DMG (Table 19). However, the immediately-released bioavailable nutrients from the dredging in the Port Melbourne Channel (0.1-0.3 mg N m -2 d -1) do not make much of an addition to the existing nutrient supply from the benthos. The load from dredging is also low compared to the load carried by the Yarra (~ 1,400-2,200 t N) which may impact this area. The production of labile nutrients is certain, but the likelihood that this leads to an increase in plankton production and deposition, and then a decline in denitrification is unlikely to possible. If it did, the consequence is negligible, unlikely to be detectable from background (except perhaps directly adjacent to the channel), with recovery within a year. Note that the load is substantially lower than that previously estimated for the EES. As noted earlier, the sediments from the Port Melbourne Channel are clayey, with relatively low nutrient content. Furthermore, because it is clayey, dredge material often forms large balls, rather than becoming dispersed in the suction water from the TSHD. The balls retain their interstitial waters, and transfer direct to the PoM DMG. Any stimulation of growth, if it occurs, may be at the fringes of the plume. The risk quotient is very low. Similarly, the dredged material contains nutrients bound to particles, some of which may be released over time as particles settle. On a per unit area basis, the additional N load around the Port Melbourne Channel is trivial compared to the existing benthic DIN benthic flux (Table 20). Though the likelihood that labile nutrients will be released from particles is almost certain (30%), the amount that is released is so low that a decline in denitrification is unlikely at best, and the consequence negligible considering the spatial impact. This leads to a risk quotient of very low. In contrast, the combined loading at the PoM DMG from disposal of contaminated sediments from the Yarra, and uncontaminated sediments from the Yarra, Williamstown Channel and Port Melbourne Channel, is much larger than the existing benthic nutrient flux (Table 21). Since the PoM DMG is in an area remote from other inputs, it is also much larger than any other external load. The question is, whether this large increase in nutrient load leads to increased primary production (and, in turn, whether this affects denitrification). Hale (2006a) measured statistically significant increases (p<0.05) above background in dissolved and total N concentrations at the POMDMG, but not at the Port Melbourne Channel. Chlorophyll was not measured, so it is not clear whether the increased nutrients led to increases in phytoplankton biomass. Enesar (2006) did not observe a statistically significant increase (p>0.05) in chlorophyll biomass at a number of sites impacted by the dredge plume during or two weeks after the Trial Dredge Program, despite about 5% of the capital dredging volume being mobilised from the Port Melbourne Channel. Two

SEES Head Technical report: Nutrient Cycling 91

reasons could be advanced for this. The first is that so small an amount of immediately available N (5% of 1.5 -4.2 t; Table 17) would have been released at the dredge site that any increase in biomass would have been undetectable with routine laboratory methods. The second is that a much larger release (5% of 33-77 t; Table 18) may have happened over a number of days, but that would have been almost exclusively at the PoM DMG. Most may have been buried rather than dispersed into the water column, and if an increase in chlorophyll biomass did occur, it could have happened after dredging ceased (ie. if it arose from the delayed release over 10 or more days). However, Enesar did not have a site near the PoM DMG, and the possibility exists that phytoplankton biomass did increase, but no one observed it. Continuous monitoring of chlorophyll at a site about 5 km south of the PoM DMG (Longmore, unpublished) did not detect an increase in chlorophyll at the surface during or over the month after dredging. The predicted DIN loading at the DMG from particle nutrient release from the Port Melbourne Channel sediments is similar to the loadings from the Yarra River and Williamstown Channel (Table 20). Assuming the Port Melbourne Channel sediments dredged in 2005 had similar nutrient release properties to those determined in 2006, then the loading on the PoM DMG during the Trial Dredge program would have been similar to that expected for the CDP. Though the period of impact will be much longer during the CDP, the expectation of a detectable increase in phytoplankton biomass over background variation at the PoM DMG is not great. Further, benthic flux measurements on the PoM DMG in February 2006, north of the area of placement, showed very high denitrification efficiency (93 ± 2 %; Longmore unpublished), indicating that if placement in September 2005 did have an impact on denitrification at the PoM DMG, it was no longer measurable five months later. Whether there is an impact on denitrification on the PoM DMG or not, the area of the PoM DMG is small in relation to the rest of the project area, so that any impact will be small on a regional basis. For these reasons, the likelihood of algal blooms leading to a decline in denitrification is classed as possible, and the consequence to the project area is minor. The risk quotient is very low. The TSHD dredge plume will mobilise algal cysts (this has already been observed). However, the occurrence of cysts is thought to be much less in the Williamstown Channel than further north. There is no certainty that this will lead to an algal bloom, but such an outcome is possible, and is probably higher in summer-autumn than other seasons. The main impacts are: (i) economic (transfer of a toxic bloom to an aquaculture site) but the only site adjacent is not in commercial production, and (ii) ecological, on benthic organisms and nutrient cycling, if the bloom is large enough to consume bottom DO on settling. We estimate a 30% chance of that happening, so that the net likelihood is unlikely to possible. The consequence is minor, because such a bloom is unlikely to impact a significant fraction of this large project area. For example, denitrification efficiency may be reduced in the in the area around the Port Melbourne Channel, but will recover within a year. Likewise, a bloom on the PoM DMG, if it were restricted to that area, would be a negligible to minor consequence (in terms of nutrient cycling). The risk quotient is very low. Blooms of toxic algae have only been observed twice at Beaumaris and Sandringham, both within the near-shore area (see Appendix 1). While there were no parallel observations in deeper waters, the assumption is made here that such blooms have been restricted to the coast. Similar impacts, at smaller scales, were assessed for such activities as placement of the bund and placement of material to cap the contaminated sediments. They have not been listed here; all risk quotients were very low. The impact of turbidity on phytoplankton productivity, and then on denitrification, is uncertain. As mentioned for the Yarra and Hobsons Bay area, there may be a range of impacts across the area on phytoplankton growth, ranging from nutrient stimulation to turbidity suppression. Turbidity is greatest along and east of the Port Melbourne Channel, and at the PoM DMG, but has some impact over about a fifth of the project area (Fig 43). There is no doubt that phytoplankton productivity will be suppressed near the Port Melbourne Channel, both because the turbidity is high, and because the bioavailable nutrient release is low in this area. Denitrification efficiency is already high east of the Port Melbourne Channel, and supply of organic matter from the Yarra River- Hobsons Bay area may compensate for any local decline in deposition, so that the chance of an impact on denitrification is unlikely. Because recovery of phytoplankton production should be rapid after dredging, and because the area of suppression is low, the significance of any impact on denitrification in the regional context is minor at most. The risk quotient is very low.

SEES Head Technical Report: Nutrient Cycling 92

Creation of the dredge plume will lead to shading of the benthos, including the MPB. The MPB may respond with a decline in biomass or productivity, or both. Areas of particularly high MPB biomass have been observed on either side of the Port Melbourne Channel, and north of the PoM DMG (Edmunds et al . 2005). The area north of the PoM DMG is unlikely to be impacted by a plume. In the risk assessment, we have said that an impact on MPB is certain, although the experimental studies described in Chapter 6 indicate this may not be easy to distinguish from background variation. However, the experimental evidence also suggests that the likelihood that such an impact then leads to a significant decline in denitrification is possible rather than certain. If we assume denitrification flux in the area affected by the plume is 25 mg N m -2 d -1 (Table 10), and MPB contribute to 50% of this, the loss of denitrification under the plume may be ~ 12 mg N m -2 d -1, which exceeds the current DIN flux of 2-7 mg N m -2 d -1 (Table 10). However, because the plume is expected to affect less than 20% of the project area (Fig 43), the impact on the project area as a whole is 2-3 mg N m -2 d -1, within the range of natural variation. The consequence is rated as minor, since the impact is likely to be restricted to areas under the plume (around the Port Melbourne Channel), and recovery is likely within a year. The risk quotient is very low. The area impacted by sedimentation (not shown) is similar to that affected by the 2 mg L -1 contour. Sedimentation from the dredge plume may impact on MPB. There is evidence that some types of cells in the MPB can migrate in the sediment to compensate for burial (Edmunds et al . 2004), and given the spatial distribution of MPB biomass, the areas most vulnerable are patches of high biomass either side of the Port Melbourne Channel. We have already argued that the evidence for a significant role of MPB in denitrification is thin. Even so, in a worst case, we can assume that all MPB productivity is lost from the 50 km 2 area on and east of the Port Melbourne Channel, and that MPB makes a 50% contribution to denitrification. By a similar calculation as described for this impact in Hobsons Bay, we can estimate the loss of denitrification as a result of the loss of MPB productivity, as the equivalent of an increased N load of about 7 tonnes N. In a regional context, this is a greater load than that from release of immediately available N in the plume (Table 16), but still only 10-25% of the existing benthic DIN flux. Even if recovery took a year, for the same reasons listed above, with a likelihood of possible, and a consequence of minor, the risk quotient is very low. Sedimentation from the dredge plume may lead to the burial of infauna, and therefore of whatever contribution they make to denitrification. Modelling (not shown here) indicates heavy sedimentation on either side of the Port Melbourne Channel (and at the PoM DMG). However, for most of the management area, the sedimentation is expected to be very light (< 100 g m -2). Sediment traps in central Port Phillip Bay measured deposition rates of about 0.1-0.2 g m -2 h -1 (You et al . 1996), suggesting that natural sedimentation over a 16-week period similar to the dredging (~300-600 g m -2) exceeds that expected from dredging. The existing fauna may therefore be adapted to such deposition rates, and the extra from dredging may make little difference to faunal activity. Similarly, the fauna in the PoM DMG have undoubtedly been impacted many times in the past, due to maintenance dredging. This may not be true for those areas subject to heavier sedimentation, but they comprise less than a fifth of the total project area. On this basis, though sedimentation is certain, the probability that it leads to a loss in infauna is rated as possible, and while the consequence in those areas affected may be moderate, the consequence to the region is minor, with recovery expected within a year. The risk quotient is therefore very low. Toxicants may affect grazers on phytoplankton. However, none of the Port Melbourne Channel sediments are contaminated, so that this will not be an issue in the plume. Placement of contaminated sediment from the Yarra will be by diffuser close to the PoM DMG floor, avoiding the creation of a plume, and restricting any impact to the PoM DMG. Any toxicants in the dissolved phase are likely to be rapidly sorbed to particles, and re-deposited on the Bay floor. The likelihood of an impact on zooplankton, even at the PoM DMG, is therefore possible at best. If there was an impact, it will be confined to the PoM DMG. The likelihood of being able to detect the flow-on impact on denitrification, especially given the massive deposition of sediments at the same time, is low. Given the restricted area of impact, the consequence is negligible, and the risk quotient is very low. Toxicants may affect infauna, but only in the PoM DMG. Even if we “write off” the PoM DMG as an area contributing to denitrification, the area of the PoM DMG is so small in the regional context, that this will make no practical difference to the capacity of the region to denitrify. The likelihood of an impact is possible; the consequence is minor, and the risk quotient is very low. Note that simple burial of the sediment surface by uncontaminated dredge material placement will have the same effect, and it would be impossible to separate the effect on denitrification from either of these impacts.

SEES Head Technical report: Nutrient Cycling 93

Table 27 Summary of risks to nutrient cycling in the north of the Bay.

Predicted effects Aspect Activity Initiating Impact A Impact B Consequence event level Dredging Phsical Removal of Loss of Negligible TSHD removal of seabed effective seabed denitrification layer Risk events Aspect Activity Initiating Impact A Impact B Conse- Risk event quence Quotient level Dredging Dredging - Increased Algal blooms Minor Very Low TSHD creation of bioavailable (non-toxic) plume nutrients Dredging Dredging - Mobilisation Algal blooms Transfer to Minor Very Low TSHD creation of of algal cysts (toxic) sensitive plume area Dredging Dredging - Increased Increased Algal Minor Very Low TSHD creation of suspended bioavailable blooms (non- plume sediments nutrients toxic) Dredging Dredging - Increased Increased Reduction in Minor Very Low TSHD creation of suspended bioavailable denitrificatio plume sediments nutrients n Dredging Dredging - Increased Increased Reduction in Minor Very Low TSHD creation of suspended bioavailable denitrificatio plume sediments nutrients n Bunded Placement of Increased Reduction in Minor Very Low DMG capping bioavailable denitrification material - nutrients creation of plume Dredging Transit of Translocation Denitrification Moderate Very Low TSHD dredging of marine change due to vessel / pests altered fauna barge to / from DMG (loaded / unloaded) Bunded Placement of Translocation Denitrification Minor Very Low DMG material of marine change due to within pests to DMG altered fauna bunded area

SEES Head Technical Report: Nutrient Cycling 94

- marine pests

There is the possibility that marine pests may be translocated during transit from the Yarra/Williamstown dredging area to the PoM DMG. The chance that this would occur, and then lead to a change in denitrification, was rated as no higher than possible, because of the following reasons. Firstly, there is no evidence that there is a substantial difference now in the marine pests present in the Yarra/Williamstown area and at the PoM DMG. Secondly, even if there were, there is no evidence that transit will lead to the dispersion of such pests Thirdly, such a pest would only flourish if there was a niche at the PoM DMG different to that at the dredge area. Finally, we know of only one pest currently in the Bay ( Raeta pulchella, which is already found from Hobsons Bay to Fawkner Beacon; Ross et al . unpublished) which appears to have a significant impact on denitrification at current density. Because Raeta occurs in the area to be dredged, it could be transferred to the PoM DMG. If the new colony survived and was restricted to the PoM DMG, the impact on the PoM DMG may be significant, but the regional impact on denitrification is minor. The impacts of fauna on denitrification are often “patchy”, so that high denitrification can still be measured adjacent to the Port Melbourne Channel, even though such fauna occur north and south of the flux monitoring site. To have a significant impact on denitrification, exotic fauna probably need to colonise in a uniformly dense manner, or alter the infauna/microbial population over very wide areas. The risk quotient is very low, and again, this impact on the PoM DMG would not be distinguishable from the impacts of burial or toxicity until the microbial population recovered in the (new) surface sediments. The combined effects of labile nutrients, turbidity and smothering did not lead to a statistically significant decline in denitrification at a site 300 m east of the Port Melbourne Channel during the Trial Dredge Program (Longmore 2006b). It may be that the trial ran for too short a period to detect some of these impacts. Similarly, denitrification efficiency on the PoM DMG (north of the area of placement) had recovered within five months of the trial. Cumulative impacts on denitrification therefore did not appear to lead to significantly higher consequences than each impact considered separately.

SEES Head Technical report: Nutrient Cycling 95

7.4 South of the Bay

The two highest risks in this area are rated at low. They are the impacts of turbidity on phytoplankton and MPB. Dredging activities carried out in this management area include dredging in the South Channel, and placement at the SE DMG. Dredging will take place in three periods, of 16, 13 and 12 weeks duration, spread over 12 months. This is an area with no significant external source of nutrients, and generally low ambient nutrient and phytoplankton concentrations. Its role in the denitrification capacity of the Bay is uncertain, but likely to be higher than the CSIRO model predicted. Activities which were assessed are listed here (Table 28). Predicted events Physical removal of the seabed is certain; subsequent loss of an effective denitrification layer is less certain but probable. Though this is in an area assigned a very low denitrification capacity by the PPBES model, subsequent experimental measurements indicate this may not always be true. The consequence to the denitrification capacity of the area is considered negligible, because the channel bed comprises only a very small proportion of the seabed. The risk quotient is very low. Risk events

As noted earlier, the sediments from the South Channel are sandy, with relatively low nutrient content. Even so, the dredged material contains labile nutrients, which will mostly be released in the South Channel in overflow, with a minor portion released at the DMG (Table 19). Given the size of the load (only about 8-14 tonnes ammonium-N), the very large area over which dissolved nutrients may be dispersed (~ 600 km 2), and the period of dredging in this area (spread over two years), compared to the N flux already carried by the tidal flow (~ 0.1 mg N m -2 d -1), this load amounts to an approximate doubling over background. However, it is insignificant compared to the estimated current benthic DIN flux (Table 20). The production of labile nutrients is certain, but the likelihood that this leads to a decline in denitrification (via increased algal growth) is possible. If it did, the consequence is minor, likely to be detectable from background, with recovery within a year. Note that the load is substantially lower than that previously estimated for the EES. The risk quotient is very low. Similarly, the dredged material contains nutrients bound to particles, some of which may be released over time as particles settle. The vast majority of this release, if it occurs, will be in the SEDMG, and much may not occur, because most of the sediments will be buried, rather than exposed to the water column. The likelihood that labile nutrients will be released from particles is certain, but the amount likely to be released is low, and mostly attributed to nitrate, because several of the sediment samples from the South Channel took up ammonium over 10 days, rather than released it. Even if all the release from particles did occur, and at the SEDMG, this would be a DIN load of a similar magnitude to the existing benthic DIN flux at the SEDMG. The probability that this will lead to an increase in phytoplankton production, and then to a decline in denitrification is unlikely. However, given the small proportion of the project area likely to be affected (the SEDMG), the consequence is minor: a measurable impact, but recoverable within a year of the end of dredging. The risk quotient is very low. Enesar (2006) did not observe a significant increase in chlorophyll biomass at a number of sites impacted by the dredge plume during or two weeks after the Trial Dredge Program, despite about 5% of the capital dredging volume being mobilised from the South Channel. Two reasons could be advanced for this. The first is that so small an amount of immediately available N (5% of 7.9 ± 6.2 t; Table 17) would have been released at the dredge site and dispersed with the plume that any increase in biomass would have been undetectable with routine laboratory methods. The second is that a much larger release may have happened over a number of days (5% of 2.0 ± 40 t; Table 18), but that would have been almost exclusively at the DMG, most may have been buried rather than released, and if it occurred, may have led to an increase in biomass after the post-dredge sampling. No monitoring was carried out over the SEDMG. AS for the North above, Hale (2006a) carried out measurements in the trial dredge plume, and observed total

SEES Head Technical Report: Nutrient Cycling 96

N and nitrate concentrations statistically significantly higher than background in the South Channel. No statistically significant increase was observed at the SEDMG, and no chlorophyll measurements were made.

The TSHD dredge plume will mobilise algal cysts where present (this has already been observed). However, there are no records of toxic algal cysts in the South Channel sediments. The major concern with blooms in this area could be the enhanced possibility of bitter taste ( Rhizosolenia ) blooms. This species is known to form cysts, and given the almost total coverage of the Bay during Rhizosolenia blooms, the cysts are likely to be in the South Channel sediments. There is no certainty that suspension of the cysts will lead to an algal bloom, and such an outcome is possible (could have occurred already), and is probably higher in winter than other seasons. The main impacts are: (i) economic (transfer of toxic bloom to aquaculture sites), and (ii) ecological, on benthic organisms and nutrient cycling, if the bloom is large enough to consume bottom DO on settling. We estimate a 30% chance of that happening, so that the net likelihood is very unlikely to possible. The consequence is minor, because such a bloom is unlikely to impact a significant fraction of this large project area. Given the tidal movement, the bloom may be restricted to the area close to the channel, rather than spread across the Sands or further north. For example, denitrification efficiency may be reduced in the in the area around the South Channel, but will recover within a year. Likewise, a bloom on the DMG, if it were restricted to that area, would be a negligible to minor consequence, in terms of nutrient cycling, to the Project Area as a whole. The risk quotient is very low. Hale (2006b, Appendix 1 in this report) concluded that the balance between nutrient enhancement of phytoplankton growth, and turbidity suppression was tilted toward turbidity in the South Channel. Because primary production is already low, and the area of the plume is small compared to the Project Area, she concluded that it was unlikely there would be any long-term impact on the food chain. However, precisely because this is an area of low production, low deposition, high tidal mixing and organically-poor sediments, of all the areas of the Bay, this is the one most prone to a decrease in denitrification due to oxygen inhibition, as sediments become more oxic. The CSIRO PPBES model predicted low denitrification efficiency on and south of the Sands for this reason. This may be already happening naturally on occasion, and could explain the quite large changes in denitrification efficiency over months that have been observed on the Sands. The probability of turbidity leading to a decline in denitrification is therefore highly probable (it may have already occurred here), but because of the likelihood of rapid recovery once dredging ceases, the consequence for the area is minor. The risk quotient is low. Creation of the dredge plume will lead to shading of the benthos, including the MPB. The northern side of the Sands is host to some of the highest MPB biomass in the Bay (Edmunds et al. 2005), much of which will be impacted by the plume for significant proportions of the time. The MPB may respond with a decline in biomass or productivity, or both. In the risk assessment, we have said that such a response is certain, although the experimental studies described in Chapter 6 indicate this may not be easy to distinguish from background variation. However, the experimental evidence also suggests that the likelihood that such an impact then leads to a significant decline in denitrification is probable (30%) rather than certain. The likelihood may be slightly higher in the south than in the north, because the water clarity is usually higher in the south than in the north, and the impact of turbidity may be proportionally greater than in the more turbid north. If MPB was responsible for 50% of the denitrification in the south, and the N 2 flux in the south averages 6 mg N m -2 d -1 (Longmore, unpublished) loss of MPB productivity could lead to an increased DIN flux of ~ 3 mg N m -2 d -1 under the plume. This is within the range of DIN fluxes measured in the south (Table 21). Since MPB is very patchy, and the trial dredge experiment was carried out in an area of high biomass, if an impact on MPB was not observed at that site, it seems unlikely that an impact would be detectable at sites of lower MPB biomass. The calculation above is therefore probably conservative. The consequence is rated as minor, since the impact is likely to be restricted to areas under the plume, and recovery is likely in much less than a year. The risk quotient is low. The impact of sedimentation on MPB would be similar to the impact of turbidity, but over a smaller area, albeit for a longer time since MPB would have to recolonise after burial, rather than just resume

SEES Head Technical report: Nutrient Cycling 97

photosynthesis. However, for all practical purposes this impact is already accounted for in the impact of turbidity. The risk quotient is therefore low. Much of the sediments in the South are highly mobile, and the faunal community surviving under the area likely to be affected by sedimentation will likely have adapted to heavy episodic burial. Such occasional drastic rearrangements of the faunal community may be another explanation for the apparent large changes over time in denitrification efficiency in the south. The impact of sedimentation from dredging, which is heaviest around the Hovell Pile, may not be distinguishable from the natural variation. The likelihood of an impact is possible to highly probable (it may have occurred before), but the consequence is probably negligible in a regional context (indistinguishable from background). The risk quotient is very low. Because the sediments from the South Channel are non-toxic, the issues of the impact of toxicants on grazers, and toxicants on infauna, do not arise. The sediments in the south are quite different to those in the rest of the project areas (sandy rather than muddy). The translocation of exotic species from elsewhere in the Bay is very unlikely to unlikely, and the consequence to denitrification is minor to moderate. The risk quotient is very low.

SEES Head Technical Report: Nutrient Cycling 98

Table 28 Summary of risks to nutrient cycling in the south of the Bay. Predicted effects Aspect Activity Initiating Impact A Impact B Consequence event level Dredging Dredging - Increased Reduced light Impacts Minor TSHD creation of suspended photosynthesis plume sediments (MPB/phyto) Non-bunded Placement of Smothering of Reduction in Minor DMG material sea floor inside denitrification DMG Risk events Aspect Activity Initiating Impact A Impact B Conse Risk event quence Quotient level Dredging Dredging - Increased Reduction in Minor Very Low TSHD creation of bioavailable denitrification plume nutrients Dredging Dredging - Mobilisation Algal blooms Transfer to Minor Very Low TSHD creation of of algal cysts sensitive area plume Dredging Dredging - Increased Increased Algal blooms Minor Very Low TSHD creation of suspended bioavailable plume sediments nutrients Non- Placement of Increased Reduction in Minor Very Low bunded material - bioavailable denitrification DMG creation of nutrients plume Non- Placement of Increased Increased Algal blooms Minor Very Low bunded material - suspended bioavailable DMG creation of sediments nutrients plume Dredging Dredging - Increased Increased Reduction in Minor Very Low TSHD creation of suspended bioavailable denitrification plume sediments nutrients

SEES Head Technical report: Nutrient Cycling 99

7.5 The Entrance

All risks to nutrient cycling were rated very low or less in this area.

Predicted events

There are no predicted events of relevance to nutrient cycling in this area.

Risk events . The issues of turbidity, sedimentation and toxicity on denitrification in the Entrance were all assessed as less than very low risk, both because the sediments are not toxic or not expected to settle in the Entrance, and also because the rocky surfaces in the Entrance are not thought to contribute to the denitrification capacity of the Bay.

SEES Head Technical Report: Nutrient Cycling 100

7.6 Summary

The three highest risks to nutrient cycling were all rated as ”low”. The risks included: • the release of labile nutrients from the dredge plume in the Yarra/Hobsons Bay area; • the impact of turbidity on phytoplankton production in the south; • the impact of turbidity on MPB production in the south. All of the impacts are short-term, and restricted in space to part or all of individual project areas. None are Bay-wide, or likely to last longer than a year after the end of dredging.

Cumulative impacts

Potential impacts of dredging on nutrient cycling have been assessed above as though all impacts were independent. The response of the system to two or more independent impacts may be: • additive (the impact of both together is the sum of each), • synergistic (the impact of both together is greater than the sum of each), or • antagonistic (the impact of one reduces the impact of the other). In this assessment, the same impact in different Project Areas is assumed to be additive. There is only one period (autumn 2007) when dredging is planned in two adjacent Project Areas (excluding the Entrance), so that the chance of additive impacts between adjacent areas is minimised by project design. In many cases, the impact in a specific area of one process (eg. loss of denitrification through burial of MPB) has been assumed to also include other impacts on the same process (eg. loss of denitrification through shading of MPB). Complete loss of denitrification in a specific area has been assumed to preclude further impact by other processes (except, perhaps, in the recovery period). Though the non-linear nature of the PPBES model response to very high nutrient inputs has been noted (eg. an exponential increase in ammonium released with declining denitrification efficiency), the impacts predicted for the CDP are not expected to be high enough to lead to synergistic effects. The cumulative impacts are expected to be within the region of linear response, so that treating impacts as additive is appropriate. Antagonistic impacts (eg. competing impacts on phytoplankton growth between nutrient enrichment and turbidity) have been discussed where appropriate. In each assessment, likelihood and /or consequence have been estimated conservatively to allow for cumulative impacts from unassessed or unknown processes. Unless the cumulative impact increases the risk by a factor of 10, the risk quotient should be unaffected. Furthermore, the group workshops that assessed risk involved discussion in the context of the overall project activities, with much discussion between different disciplines to ensure that combined effects were assessed wherever possible. Risks to nutrient cycling from dredging were also assessed against the background of known inputs. With our current knowledge of the area and duration of impacts from known events (eg. floods), the additional risk from dredging has been assessed as increasing risk by less than an order of magnitude. It could be argued that all of the measurements of denitrification in the Bay have been collected during a period of relatively low runoff, and this analysis may underestimate the impact of flood events (and hence underestimate the cumulative impact of dredging on top of natural events). However, Parslow et al. (2000) found that the three highest 30-day peak flows in the Yarra since 1976 occurred during the PPBES baseline on which the PPBES model was based. The last of these, in September 1993, occurred 15 months before the first measurements of denitrification were made. While we are reliant on the model to predict what denitrification may have been in response to these floods, it is clear from the measurements that denitrification efficiency 15 months after the floods was similar to recent measurements. This indicates that the system had recovered from any impact within 15 months. On the basis of the scale used in this analysis (Table 22), the consequence of such a flood to nutrient cycling was moderate at most.

SEES Head Technical report: Nutrient Cycling 101

Uncertainty in relation to risk assessment

No major risk to nutrient cycling, and denitrification in particular, has been identified from our current knowledge. While the use of denitrification efficiency as an indicator of Bay health is soundly based on observations and modelling, there are many aspects of denitrification that are still unclear. These include: • The reason(s) for quite large changes in efficiency at some sites, in both the north and south, over a period of months, with apparently full recovery; • Whether dredging will exacerbate such changes; • Whether the failure to define a strong link between MPB and denitrification during the Trial Dredge Program means that MPB over the Bay as a whole can be ignored as far as nutrient cycling is concerned.

Because the spatial and temporal variability of many of the factors affecting nutrient cycling is unknown, there is a degree of uncertainty in the risk assessment carried out. The factor of 10 intervals in estimates of likelihood and consequence described above, allow for a wide range in uncertainty about the magnitude of various impacts on nutrient cycling. Impacts relating to loss of denitrification surface in the channels must be small, because of the relatively small area affected. Impacts relating to particle transport, turbidity or burial are constrained by the size of the dredge plume and/or area of DMG. In this assessment, a conservative approach has been taken to the definition of plume size (using the greatest extent of the modelled 2 mg L -1 plume), and the mean spatial impact may be much smaller than estimated here. Impacts relating to phytoplankton growth may extend beyond the plume, but the additional biomass is expected to be within natural variation. The risk of an unexpected impact on nutrient cycling from this route is therefore low. The largest remaining uncertainties include the impact of dredging on denitrification via : impacts on infauna; impacts on MPB, and the potential for a bay-wide invasion of exotic marine pests. A further level of uncertainty exists for those calculations based on predicted dredging volumes; there is a 15% uncertainty in the volume to be dredged.

The cumulative impacts of dredging on nutrient cycling, as far as they can be assessed, are summarised below (Table 30), in equivalents of new nitrogen inputs (Longmore 2006c). Though there is much higher confidence in some of the estimates than others, the mean impact of channel deepening on nutrient cycling is estimated here to be less than 300 t N.

SEES Head Technical Report: Nutrient Cycling 102

Table 29 Cumulative effects of impacts of dredging on nutrient cycling (Longmore 2004; Longmore 2006c).

Area Source Mean impact (t N) 95 percentile (t N) Confidence in estimate Yarra/Hobsons Immediate DIN 34 129 High Bay release Slow DIN release 45 125 High Impact via MPB 5 (5) Moderate Impact via infauna 18 (25) Low

North (incl PoM Immediate DIN 3 6 High DMG) release Slow DIN release 34 77 High Impact via MPB 40 (90) Moderate Impact via infauna 70 (90) Low

South Immediate DIN 10 16 High release Slow DIN release -3 22 High Impact via MPB 16 (32) Moderate Impact via infauna 23 (24) Low

Baywide total 285 641

Note: Numbers shown in brackets are estimates of maximum impact, rather than 95 th percentiles (Longmore 2004).

SEES Head Technical report: Nutrient Cycling 103

8 Mitigation

Because none of the risks has been assessed as moderate or higher, no specific mitigation is proposed for nutrient cycling. Many of the proposed project delivery standards will act to minimise the impact on nutrient cycling. For example, restricting dredging to defined construction zones, controls on placement at the DMGs, use of the green valve, controls related to limiting turbidity (eg. interval dredging, overflow/non-overflow, choices between smaller and larger dredges), selection of dredging periods in north and south, may all reduce the effect of dredging on nutrient cycling. Project Delivery Standards (referred to as threshold limits and performance criteria in the SEES Assessment Guidelines) will be addressed in a separate document (the project Environmental Management Plan). These include a strategy to manage the risk of algal blooms. Given the uncertainties surrounding the impact of dredging on nutrient cycling described above, it would be prudent to monitor denitrification efficiency at a number of key sites throughout the dredging program, and for some period afterward. The method, rationale and statistical basis for this will also be described in the Environmental Management Plan. No recommendations are made for offsets to the N load likely to be introduced by dredging, because mitigation is not needed.

SEES Head Technical Report: Nutrient Cycling 104

9 Conclusions

Port Phillip Bay appears to be unusually efficient at denitrification, and denitrification has been identified as a key indicator to be maintained. While denitrification is efficient over much of the Bay floor, there is high spatial and moderate temporal variability. Efficiency is lower in areas close to the major inputs (Hobsons Bay and the Werribee coast) than in the centre of the Bay. There is no evidence that average annual efficiency has changed by more than about 10% over the past 10 years at the two long-term monitoring sites in Hobsons Bay and central Port Phillip Bay. Experiments have demonstrated that a number of species of exotic infauna already present in Port Phillip Bay may affect nutrient cycling and denitrification, but only at densities much higher than found in Port Phillip Bay. Microphytobenthos actively take up nutrients from the sediment, and produce oxygen during daylight, which may promote nitrification, but inhibits denitrification. Studies during the Trial Dredge Program did not provide compelling evidence that microphytobenthos plays a significant role in denitrification in the areas examined. Nutrient cycling, and denitrification in particular, may be vulnerable to a number of effects of dredging. Dredging and subsequent dredged material placement has the potential to affect benthic organisms, ambient light, nutrient and oxygen concentrations and phytoplankton growth, all of which may influence nutrient cycling and denitrification. A formal assessment of the risk of dredging to algal blooms, nutrient cycling and denitrification found that only one bay-wide impact was thought possible. The introduction an exotic species which then caused a bay-wide decline in denitrification, while of major consequence, had such a low likelihood that the risk was extremely low. All other impacts applied to one or more of the project areas, rather than the whole of the Bay. The three highest risks to nutrient cycling were all rated at ”low”. The risks included: • the release of labile nutrients from the dredge plume in the Yarra/Hobsons Bay area; • the impact of turbidity on phytoplankton production in the south; • the impact of turbidity on MPB production in the south. All of the impacts are short-term, and restricted in space to part or all of individual project areas. None are Bay-wide, or likely to last longer than a year after the end of dredging. No mitigation is proposed specifically for nutrient cycling, but it is recognized that many of the project design standards will act to minimize the effect of dredging on nutrient cycling. Even so, because of the continuing uncertainty about specific mechanisms affecting denitrification, and the potentially disastrous impact on the Bay if denitrification fails, monitoring is proposed before, during and after dredging to confirm that high denitrification efficiency is maintained throughout the Channel Deepening Project.

SEES Head Technical report: Nutrient Cycling 105

10 References

Aller RC, Aller JY 1992. Meiofauna and solute transport in marine muds. Limnology and Oceanography 37 , 1018-1033. ANZECC/ARMCANZ 2000. Australian and New Zealand Guidelines for Fresh and Marine Water Quality. Australian and New Zealand Environment and Conservation Council/Agriculture and Resource Management Council of Australia and New Zealand. Arnott GH, Conron SD, Reilly DJ, Hill DRA, Sonneman JA 1994. Effects of channel-maintenance dredging on the mobilization and translocation of toxic algal cysts in Port Phillip Bay. Report to the Port of Melbourne Authority. Arnott GH, Hill DRA, Reilly DJ, Gason AS, Magro KL, Coots AG 1997. The composition and distribution of phytoplankton in Port Phillip Bay from March 1990 to February 1995. Technical report No. 40, CSIRO Port Phillip Bay Environmental Study, ACT. Beardall J, Roberts S, Royle R 1997. Phytoplankton productivity in Port Phillip Bay: seasonal and spatial distributions. Technical Report No. 35, CSIRO Port Phillip Bay Environmental Study, ACT. Beattie G, Redden A, Royle R 1997. Microzooplankton grazing in phytoplankton in Port Phillip Bay. Technical Report No. 37, CSIRO Port Phillip Bay Environmental Study, ACT. Berelson WM, Heggie D, Longmore A, Kilgore T, Nicholson G, Skyring G 1998. Benthic nutrient recycling in Port Phillip Bay, Australia. Estuarine, Coastal and Shelf Science 46, 917-934. Berelson WM, Townsend T, Heggie DT, Ford P, Longmore AR, Skyring GW, Kilgore T, Nicholson GJ 1999. Modelling bio-irrigation rates in the sediments of Port Phillip Bay. Marine and Freshwater Research 50, 573-580. Bird FL, Ford P, Heislers S 1996. Bioturbation by common Port Phillip Bay benthic invertebrates. Technical Report No. 37, CSIRO Port Phillip Bay Environmental Study, ACT. Bird FL, Ford FW, Hancock GW 1999. Effect of burrowing macrobenthos on the flux of dissolved substances across the water-sediment interface. Marine and Freshwater Research 50, 523-532. Burke CM 1999. Molecular diffusive fluxes of oxygen in sediments of Port Phillip Bay in south-eastern Australia. Mar. Freshwater Res . 50 , 557-566. Devol AH 2003. Nitrogen cycle- Solution to a marine mystery. Nature 422, 575-576. DNRE 2002. Port Phillip Bay Environmental Management Plan: Plan and Critical Programs to 2003. Department of Natural Resources and Environment, Melbourne. DSE 2005. Assessment guidelines issued by the Minister for Planning for Port Phillip Bay Channel Deepening Supplementary Environment Effects Statement. Department of Sustainability and Environment. Edmunds M, Beardall J, Hart S, Elias J, Stojkovic-Tadic S 2004. Port Phillip Bay Channel Deepening Project Environment Effects Statement- Marine Ecology Specialist Studies. Volume 4- Microphytobenthos. Australian Marine Ecology Report 161, Melbourne, 45 pp. Edmunds M, Bryant C, Pickett P, Crozier J 2005. Port Phillip Bay Channel Deepening Project Trial Dredge Experiment: Microphytobenthos. Report No. 332, Australian Marine Ecology, Melbourne. Enesar 2006. Trial dredge program Mobile Water Quality Monitoring- Program X4. Enesar P/L, Melbourne. Engstrom P, Dalsgaard T, Hulth S, Aller RC 2005. Anaerobic ammonium oxidation by nitrite (anammox): Implications for N2 production in coastal marine sediments. Geocimica et Cosmochimica Acta 69, 2057-65.

SEES Head Technical Report: Nutrient Cycling 106

EPA 1992. Trial Dredge Protocol. Environment Protection Authority, Victoria, Publication 312. EPA 1997. Protecting Water Quality in Port Phillip Bay, SEPP (Waters of Victoria) Schedule F6. Environment Protection Authority, Victoria, Publication 434. EPA 2001. Best Practice Environmental Management Guidelines for Dredging. Environment Protection Authority, Victoria, Publication 691. EPA 2002. Environmental Report. Port Phillip Bay Water Quality. Long term trends in nutrient status and water clarity, 1984-99. Publication 806, Environment Protection Authority, Victoria. Fletcher TD, Deletic A 2006. A review of Melbourne Water’s pollutant load monitoring program for Port Phillip and Western Port. Benchmark Environmental Consulting, 86 p. Gorfine HK, Longmore AR, Abbott BD 1996. Sediment oxygen demand in Port Phillip Bay. Publication 495, Environment Protection Authority, Victoria. Grant BR, Gayler KR 1982. Phytoplankton productivity study. Supplementary Paper 6. Pp 255-326 in: The Heated Effluent Study - Final Report . (Eds: Sandiford RN, Holmes N). Marine Science Laboratories Technical Report No. 15, Marine Science Laboratories, Victoria. Hale J 2006a. Supplementary Environment Effects Statement, Trial Dredge Program. X1.1 Nutrient and Contaminant Monitoring in the Dredge Plume. Port of Melbourne Corporation. Hale J 2006b. Channel Deepening Project: Potential Effects on Phytoplankton. Report appended to this report. Harris G, Batley G, Fox G, Hall D, Jernakoff P, Molloy R, Murray A, Newell B, Parslow J, Skyring G, Walker S 1996. Port Phillip Bay Environmental Study Final Report. CSIRO, ACT. Heggie DT, Skyring GW, Orchardo JW, Berelson WM, Longmore AR, Nicholson GJ 1999. Denitrification and denitrifying efficiencies in sediments of Port Phillip Bay: direct determinations of biogenic N 2 and N- metabolite fluxes with implications for water quality. Marine and Freshwater Research 50, 589-596. Irigoien X, Castel J 1997. Light limitation and distribution of pigments in a highly turbid estuary: the Gironde (SW France). Estuarine, Coastal and Shelf Science 44, 507-517. Kemp WM, Boynton WR, Adolf JE, Boesch DF, Boicourt WC, Brush G, Cornwell JC, Fisher TR, Glibert PM, Hagy JD, Harding LW, Houde ED, Kimmel DG, Miller WD, Newell RIE, Roman MR, Smith EM, Stevenson JC 2005. Eutrophication of Chesapeake Bay: historical trends and ecological interactions. Marine Ecology Progress Series 303, 1-29. Kowarsky J 2002. South Channel Eastern Entrance Port Melbourne Channel Beacons 1 to 6 - Proposed Dredging 2002 Sediment Quality Report. Report for Sinclair Knight Merz by John Kowarsky and Associates. Longmore AR 1992. Nutrients in Port Phillip Bay: what has changed in 15 years? Proc. Roy. Soc. Vic . 104, 67-73. Longmore AR 2004. Channel Deepening Project Proof of Concept Modelling. Revised calculations of nitrogen cycling and denitrification impacts. Document tendered to the Panel. Longmore AR 2005. Port Phillip Bay Environmental Management Plan: Monitoring the state of bay nitrogen cycling (2002-2005). Marine and Freshwater Systems Report No. 7, Primary Industries Research Victoria, Queenscliff. Longmore AR 2006a. Supplementary Environment Effects Statement, Trial dredge program: X9 (Sediment Nitrogen). Marine and Freshwater Systems Report No. 11, Primary Industries Research Victoria, Queenscliff. Longmore AR 2006b. Supplementary Environment Effects Statement, Trial dredge program: X8 (MPB- denitrification). Marine and Freshwater Systems Report No. 10, Primary Industries Research Victoria, Queenscliff. Longmore AR 2006c. Supplementary Environment Effects Statement: Sediment Nutrient Availability. Marine and Freshwater Systems Report No. 15, Primary Industries Research Victoria, Queenscliff.

SEES Head Technical report: Nutrient Cycling 107

Longmore AR 2006d. Channel Deepening Project: Further study of the relationships between suspended solids, turbidity and light attenuation. Marine and Freshwater Systems Report No. 16, Primary Industries Research Victoria, Queenscliff Longmore AR, Cowdell RA, Flint R 1996. Nutrient status of the water in Port Phillip Bay. Technical Report No. 24, CSIRO Port Phillip Bay Environmental Study, ACT. Longmore AR, Cowdell RA, Abbott B, Sporcic M 1997. Water quality in Port Phillip Bay 1997: Final Report for the Victorian Fisheries Port Phillip Bay Environmental Monitoring Program. Marine and Freshwater Resources Institute Report Technical Report, Queenscliff. Longmore A, Nicholson G, Parry G 1996a. The impact of Sabella spallanzanii on infauna and nitrogen cycling in Port Phillip Bay. Marine and Freshwater Resources Institute Report Technical Report No. 4, Queenscliff. Longmore AR, Gason A 2001. Port Phillip Bay Nitrogen Monitoring Proposal- Statistical Advice. Marine and Freshwater Resources Institute Report No. 32, Queenscliff Victoria. Longmore AR, Brady B, Fabris GJ 2004. Port Phillip Bay Channel Deepening Environment Effects Statement: Volume 2- Risk Report, Sediment Chemistry and Water Quality Specialist Study. Internal Report No. 21, Marine and Freshwater Systems, Primary Industries research Victoria, Queenscliff. Macreadie PI 2004. Comparing benthic chambers and sediment cores for measuring denitrification. Honours thesis, Zoology Department, University of Melbourne, Parkville. Macreadie P I, Ross D J, Longmore AR, Keough M J 2006. Measuring sediment denitrification. Marine Ecology Progress Series (in press; accepted April 2006). Mayer MS, Schaffner L, Kemp WM 1995. Nitrification potentials of benthic macrofaunal tubes and burrow walls; effects of sediment NH 4 and animal irrigation behaviour. Marine Ecology Progress Series 121 , 157-169. Muller FU, Larsen A, Stedman CA, Soendergaard M 2005. Interactions between algal-bacterial populations and trace metals in fjord surface waters during a nutrient-stimulated summer bloom. Limnol. Oceanogr. 50, 1855-1871. Murray A 1994. Western Treatment Plant outputs to Port Phillip Bay. Technical Report No. 15, CSIRO Port Phillip Bay Environmental Study, ACT. Murray A and Parslow J 1997. Port Phillip Bay Integrated Model: final report. Technical Report No. 44, CSIRO Port Phillip Bay Environmental Study, ACT. Murray A and Parslow J 1999. Modelling of nutrient impacts in Port Phillip Bay- a semi-enclosed marine Australian ecosystem. Mar. Freshwater Res . 50 , 597-611. Nayar S, Goh BPL, Chou LM 2004. Environmental impact of heavy metals from dredged and resuspended sediments on phytoplankton and bacteria assessed in in situ microcosms. Ecotoxicology and Environ. Safety 59, 349-369. Nicholson GJ, Longmore AR 1999. Causes of observed temporal variability of nutrient fluxes from a southern Australian marine embayment. Marine and Freshwater Research 50, 581-90 Nicholson GJ, Longmore AR, Cowdell RA 1996. Nutrient status of the sediments of Port Phillip Bay. Technical Report No. 26, CSIRO Port Phillip Bay Environmental Study, ACT. Nicholson GJ, Arnott GH, Longmore AR, Sporcic MI 1998. Dynamics of harmful Rhizosolenia cf. Chunii blooms in Port Phillip Bay. FRDC Project No. 96/264. Omnes P, Slawyk G, Garcia N, Bonin P 1996. Evidence of denitrification and nitrate ammonification in the River Rhone plume (northwestern Mediterranean Sea). Marine Ecology Progress Series 141 , 275-281. Parry GD, Langdon JS, Huisman JM 1989. Toxic effects of a bloom of the diatom Rhizosolenia chunii on shellfish in Port Phillip Bay, southeastern Australia. Marine Biology 102 , 25-41. Parslow JS, Murray A 1999. Introduced marine species and nitrogen cycling in Port Phillip Bay. Pp 275- 282. In: Hewitt CL, Campbell ML, Thresher RE and Martin RB (eds.). Marine Biological Invasions of Port

SEES Head Technical Report: Nutrient Cycling 108

Phillip Bay, Victoria. Centre for Research on Introduced Marine Pests. Technical Report No. 20, CSIRO Marine Research, Hobart. 344pp. Parslow J, Murray A, Andrewartha J, Sakov P 1999. Port Phillip Bay integrated model scenarios of nitrogen load reductions and aquaculture loads. Final Report to Victorian NRE. CSIRO, Hobart. Parslow J, Waring J, Sakov P, Sokolov S, Andrewartha J 2000. Port Phillip Bay Integrated Model Scenarios for Yarra Floods. Draft Final Report, CSIRO, Hobart. Parsons ML, Dortch Q, turner RE 2002. Sedimentological evidence of an increase in Pseudo-nitzschi (Bacillariophyceae) abundance in response to coastal eutrophication. Limnol. Oceanogr. 47, 551-558. Perez P, Estevez-Blanco P, Beiras R, Fernandez E 2006. Effect of copper on the photochemical efficiency, growth and chlorophyll A biomass of natural phytoplankton populations. Environ. Toxicol. Chem. 25, 137-143. Prytz A, Heron, ML 1999. On the flushing of Port Phillip Bay: an application of HF ocean radar . Mar. Freshwater Res . 50, 483-92. Risgaard-Petersen N, Meyer RL, Revsbech NP 2005. Denitrification and anaerobic oxidation in sediments: effects of microphytobenthos and nitrate. Aquatic Microbial Ecology 40, 67-76. Rysgaard S, Glud RN, Risgaard-Petersen N, Dalsgaard T 2004. Denitrification and anammox activity in Arctic marine sediments. Limnology and Oceanography 49, 1493-1502. Shao Q, Fox DR 1996. Spatial interpolation of underway nutrient data. Technical report No. 42, CSIRO Port Phillip Bay Environmental Study, Canberra. SKM 2004. PPB Channel Deepening EES Volume 1-Existing Conditions Report: Geological, Geotechnical and Hydrogeological Specialist Study. Sokolov S 1997. “Inputs from the Yarra River and Pattersons River/Mordialloc Main Drain into Port Phillip Bay”. Technical Report No. 33, CSIRO Port Phillip Bay Environmental Study, Canberra, Australia. Sokolov S, Black KP 1999. Long-term prediction of water quality for three types of catchment. Marine and Freshwater Research 50, 493-502. State Environment Protection Policy (SEPP) 1997. Variation of the State environment protection policy (Waters of Victoria) - insertion of Schedule F6. Waters of Port Phillip Bay. Victoria Government Gazzette No. S-101 Wednesday 27 August 1997. State Environment Protection Policy (SEPP) 1999. Variation of the State environment protection policy (Waters of Victoria) - insertion of Schedule F7. Waters of the Yarra Catchment. Victoria Government Gazette No. S-89 Tuesday 22 June 1999. Sundback K, Enoksson V, Graneli W, Pettersson K 1991. Influence of sublittoral microphytobenthos on the oxygen and nutrient flux between sediment and water: A laboratory continuous-flow study. Marine Ecology Progress Series 74 , 263-279. Sundback K, Graneli W 1988. Influence of microphytobenthos on the nutrient flux between sediment and water: a laboratory study. Marine Ecology Progress Series 43, 63-69. Swanberg IL 1991. The influenec of the filter-feeding bivalve Cerastoderma edule L. on microphytobenthos: A laboratory study. J. Exp. Mar. Biol. Ecol. 151, 93-111. Thamdrup B, Dalsgaard T 2002. Production of N2 through anaerobic ammonium oxidation coupled to nitrate reduction in marine sediments. Appl. Environ. Microbiol . 68, 1312-18. URS 2006. Draft Ecotoxicity Report. No. PR0129, 17 July 2006. You Z, Greilach PR, Turnbull JF, Hatton DN, Black KP 1996 “Field measurements of suspended sediment concentrations in Port Phillip Bay”. Technical Report No. 28, CSIRO Port Phillip Bay Environmental Study, Canberra. Walker SJ 1999. Coupled hydrodynamic and transport models of Port Phillip Bay, a semi-enclosed bay in south-eastern Australia. Mar. Freshwater Res . 50, 469-81.

SEES Head Technical report: Nutrient Cycling 109

Wood M, Beardall J 1992. “Phytoplankton ecology of Port Phillip Bay”. Technical Report No. 8, CSIRO Port Phillip Bay Environmental Study, CSIRO Special Projects Office, Melbourne.

SEES Head Technical Report: Nutrient Cycling 110

Appendix 1. Channel Deepening Project: Potential effects on Phytoplankton.

SEES Head Technical report: Nutrient Cycling 111